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Identifiable effects on public health which may be expected from the presence of a pollutant in ambient air, e.g. Heart Attacks

Submitted by Norm Roulet on Mon, 06/07/2010 - 12:00.

As a result of old science, politics and industry dominating energy, health and environmental planning and development of Cleveland, Northeast Ohio, Ohio and America, citizens here must confront the realities of too much pollution in our air today, with certainty of growing air pollution worldwide in the years ahead. As such, the United States Environmental Protection Agency's 2009 Integrated Science Assessment for Particulate Matter finds our pollution causes cardiovascular and respiratory problems and death... topping a long list of cumulative harm pollution causes people and society. Integrated Science Assessment for Particulate Matter forms the scientific foundation for the review of the primary (health-based) and secondary (welfare-based) National Ambient Air Quality Standards (NAAQS) for particulate matter (PM) in America, and "accurately reflects “the latest scientific knowledge useful in indicating the kind and extent of identifiable effects on public health which may be expected from the presence of [a] pollutant in ambient air”".

As I've long written on realNEO, Northeast Ohio has a pollution crisis and does a poor job or monitoring our pollution, putting citizens' lives in danger. How much in danger is the subject of this lengthy EPA analysis. In short, you are certainly being harmed greatly by the high levels of PM clearly released into the air in Northeast Ohio, especially near major roadways and coal burning facilities that are source points, like Mittal and MCCO. For example: "Epidemiologic studies that examined the effect of PM 2.5 on cardiovascular emergency department (ED) visits and hospital admissions reported consistent positive associations (predominantly for ischemic heart disease [IHD] and congestive heart failure [CHF]), with the majority of studies reporting increases ranging from 0.5 to 3.4% per 10 μg/m3 increase in PM 2.5".

I've summarized and included below some important highlights of this 2000+ page report... like findings related to Table 2-1 above - Summary of causal determinations for short-term exposure to PM 2.5 - concluding, among many causal things, "the collective evidence from epidemiologic, controlled human exposure, and toxicological studies is sufficient to conclude that a causal relationship exists between short-term exposures to PM 2.5 and cardiovascular effects." In other words, our air pollution is as serious as a heart attack.

One especially relevant finding I mined from page 119:

Spatial variability in source contributions across urban areas is an important consideration inassessing the likelihood of exposure error in epidemiologic studies relating health outcomes tosources. Concepts similar to those for using ambient concentrations as surrogates for personalexposures apply here. Some source attribution studies for PM 2.5 indicate that intra-urban variabilityincreases in the following order: regional sources (e.g., secondary SO 4 2– originating from EGUs)< area sources (e.g., on-road mobile sources) < point sources (e.g., metals from stacks of smelters).

Although limited information was available for PM 10-2.5 , it does indicate a similar ordering, butwithout a regional component (resulting from the short lifetime of PM 10-2.5 compared to transporttimes on the regional scale). More discussion on source contributions to PM is available inSection 3.6.

That makes living near point sources (e.g., metals from stacks of smelters) especially significant. Anyone in Cleveland live near major PM 2.5 point sources? Here are some other highlights to think about:

"The degree of spatial variability in PM was likely to be region-specific and strongly influenced by local sources and meteorological and topographic conditions (p116)." "In general, PM 2.5 has a longer atmospheric lifetime than PM 10-2.5 . As a result, PM 2.5 is more homogeneously distributed than PM 10-2.5 , whose concentrations more closely reflect proximity to local sources (Section 3.5.1.2)". "UFPs are not measured as part of AQS or any other routine regulatory network in the U.S. Therefore, information about the spatial variability of UFPs is sparse; however, their number concentrations are expected to be highly spatially and temporally variable. This has been shown on the urban scale in studies in which UFP number concentrations drop off quickly with distance from roads compared to accumulation mode particle numbers."

"Correlations between PM and gaseous copollutants, including SO 2 , NO 2 , carbon monoxide (CO) and O 3 , varied both seasonally and spatially between and within metropolitan areas (Section 3.5.3)." "The correlation between daily maximum 8-h avg O 3 and 24-h avg PM 2.5 showed the highest degree of seasonal variability with positive correlations on average in summer (avg = 0.56) and negative correlations on average in the winter (avg = -0.30). During the transition seasons, spring and fall, correlations were mixed but on average were still positive. PM 2.5 is both primary and secondary in origin, whereas O 3 is only secondary. Photochemical production of O 3 and secondary PM in the planetary boundary layer (PBL) is much slower during the winter than during other seasons. Primary pollutant concentrations (e.g., primary PM 2.5 components, NO and NO 2 ) in many urban areas are elevated in winter as the result of heating emissions, cold starts and low mixing heights. O 3 in the PBL during winter is mainly associated with air subsiding from above the boundary layer following the passage of cold fronts, and this subsiding air has much lower PM concentrations than are present in the PBL. Therefore, a negative association between O 3 and PM 2.5 is frequently observed in the winter. During summer, both O 3 and secondary PM 2.5 are produced in the PBL and in the lower free troposphere at faster rates compared to winter, and so they tend to be positively correlated."

"The federal reference methods (FRMs) for PM 2.5 and PM 10 are based on criteria outlined in the Code of Federal Regulations. They are, however, subject to several limitations that should be kept in mind when using compliance monitoring data for health studies. For example, FRM techniques are subject to the loss of semi-volatile species such as organic compounds and ammonium nitrate (especially in the West). Since FRMs based on gravimetry use 24-h integrated filter samples to collect PM mass, no information is available for variations over shorter averaging times from these instruments. However, methods have been developed to measure real-time PM mass concentrations. Real-time (or continuous and semi-continuous) measurement techniques are also available for PM species, such as particle into liquid sampler (PILS) for multiple ions analysis and aerosol mass spectrometer (AMS) for multiple components analysis (Section 3.4.1). Advances have also been achieved in PM organic speciation. New 24-h FRMs and Federal Equivalent Methods (FEMs) based on gravimetry and continuous FEMs for PM 10-2.5 are available. FRMs for PM 10-2.5 rely on calculating the difference between co-located PM 10 and PM 2.5 measurements while a dichotomous sampler is designated as an FEM."

"Results of receptor modeling calculations indicate that PM 2.5 is produced mainly by combustion of fossil fuel, either by stationary sources or by transportation. A relatively small number of broadly defined source categories, compared to the total number of chemical species that typically are measured in ambient monitoring source receptor studies, account for the majority of the observed PM mass. Some ambiguity is inherent in identifying source categories. For example, quite different mobile sources such as trucks, farm equipment, and locomotives rely on diesel engines and ancillary data is often required to resolve these sources. A compilation of study results shows that secondary SO 4 2– (derived mainly from SO 2 emitted by Electricity Generating Units [EGUs]), NO 3 – (from the oxidation of NO x emitted mainly from transportation sources and EGUs), and primary mobile source categories, constitute most of PM 2.5 (and PM 10 ) in the East. PM 10-2.5 is mainly primary in origin, having been emitted as fully formed particles derived from abrasion and crushing processes, soil disturbances, plant and insect fragments, pollens and other microorganisms, desiccation of marine aerosol emitted from bursting bubbles, and hygroscopic fine PM expanding with humidity to coarse mode. Gases such as HNO 3 can also condense directly onto preexisting coarse particles. Suspended primary coarse PM can contain Fe, Si, Al, and base cations from soil, plant and insect fragments, pollen, fungal spores, bacteria, and viruses, as well as fly ash, brake lining particles, debris, and automobile tire fragments. Quoted uncertainties in the source apportionment of constituents in ambient aerosol samples typically range from 10 to 50%. An intercomparison of source apportionment techniques indicated that the same major source categories of PM 2.5 were consistentlyidentified by several independent groups working with the same data sets. Soil-, sulfate-, residual oil-, and salt-associated mass were most clearly identified by the groups. Other sources with more ambiguous signatures, such as vegetative burning and traffic-related emissions were less consistently identified."

"Spatial variability in source contributions across urban areas is an important consideration in assessing the likelihood of exposure error in epidemiologic studies relating health outcomes to sources. Concepts similar to those for using ambient concentrations as surrogates for personal exposures apply here. Some source attribution studies for PM 2.5 indicate that intra-urban variability increases in the following order: regional sources (e.g., secondary SO 4 2– originating from EGUs) < area sources (e.g., on-road mobile sources) < point sources (e.g., metals from stacks of smelters). Although limited information was available for PM 10-2.5 , it does indicate a similar ordering, but without a regional component (resulting from the short lifetime of PM 10-2.5 compared to transport times on the regional scale). More discussion on source contributions to PM is available in Section 3.6."

In summary conclusion, Integrated Science Assessment for Particulate Matter finds, suspects and is investigating causal relationships between our air pollution and all the following concerns for our community:

2.6. Summary of Health Effects and Welfare Effects

Causal Determinations

This chapter has provided an overview of the underlying evidence used in making the causal determinations for the health and welfare effects and PM size fractions evaluated. This review builds upon the main conclusions of the last PM AQCD (U.S. EPA, 2004, 056905):

"A growing body of evidence both from epidemiological and toxicological studies supports the general conclusion that PM 2.5 (or one or more PM 2.5 components), acting alone and/or in

“A much more limited body of evidence is suggestive of associations between short-term (but not long-term) exposures to ambient coarse-fraction thoracic particles and various mortality and morbidity effects observed at times in some locations. This suggests that PM 10-2.5 , or some constituent component(s) of PM 10-2.5 , may contribute under some circumstances to increased human health risks with somewhat stronger evidence for associations with morbidity (especially respiratory) endpoints than for mortality.” (pg 9-79 and 9-80)

"Impairment of visibility in rural and urban areas is directly related to ambient concentrations of fine particles, as modulated by particle composition, size, and hygroscopic characteristics, and by relative humidity.” (pg 9-99)

“Available evidence, ranging from satellite to in situ measurements of aerosol effects on incoming solar radiation and cloud properties, is strongly indicative of an important role in climate for aerosols, but this role is still poorly quantified.” (pg 9-111)

The evaluation of the epidemiologic, toxicological, and controlled human exposure studies published since the completion of the 2004 PM AQCD have provided additional evidence for PM-related health effects. Table 2-6 provides an overview of the causal determinations for all PM size fractions and health effects. Causal determinations for PM and welfare effects, including visibility, climate, ecological effects, and materials are included in Table 2-7. Detailed discussions of the scientific evidence and rationale for these causal determinations are provided in the subsequent chapters of this ISA.

As explained in its introduction, the focus of this 2228 page Integrated Science Assessment for Particulate Matter is on scientific evidence that is most relevant to the following questions that have been taken directly from the Integrated Review Plan:

Has new information altered the body of scientific support for the occurrence of health effects following short- and/or long-term exposure to levels of fine and thoracic coarse particles found in the ambient air?

What evidence is available from recent studies focused on specific size fractions, chemical components, sources, or environments (e.g., urban and non-urban areas) of PM to inform our understanding of the nature of PM exposures that are linked to various health outcomes?

To what extent is key scientific evidence becoming available to improve our understanding of the health effects associated with various time periods of PM exposures, including not only short-term (daily or multi-day) and chronic (months to years) exposures, but also peak PM exposures (<24 hours)? To what extent is critical research becoming available that could improve our understanding of the relationship between various health endpoints and different lag periods (e.g., <1 day, single day, multi-day distributed lags)?

What data are available to improve our understanding of spatial and/or temporal heterogeneity of PM exposures considering different size fractions and/or components?

At what levels of PM exposure do health effects of concern occur? Is there evidence for the occurrence of adverse health effects at levels of PM lower than those observed previously? If so, at what levels and what are the important uncertainties associated with that evidence? What is the nature of the dose-response relationships of PM for the various health effects evaluated?

What evidence is available linking particle number concentration with adverse health effects of UF particles?

Do risk/exposure estimates suggest that exposures of concern for PM-induced health effects will occur with current ambient levels of PM or with levels that just meet the current standards? If so, are these risks/exposures of sufficient magnitude such that the health effects might reasonably be judged to be important from a public health perspective? What are the important uncertainties associated with these risk/exposure estimates?

To what extent is key evidence becoming available that could inform our understanding of subpopulations that are particularly sensitive or vulnerable to PM exposures? In the last review, sensitive or vulnerable subpopulations that appeared to be at greater risk for

PM-related effects included individuals with pre-existing heart and lung diseases, older adults, and children. Has new evidence become available to suggest additional sensitive subpopulations should be given increased focus in this review (e.g., fetuses, neonates, genetically susceptible subpopulations)?

To what extent is key evidence becoming available to inform our understanding of populations that are particularly vulnerable to PM exposures? Specifically, is there new or emerging evidence to inform our understanding of geographical, spatial, SES, and environmental justice considerations?

To what extent have important uncertainties identified in the last review been reduced and/or have new uncertainties emerged?

To what extent is new information available to inform our understanding of non-PM- exposure factors that might influence the associations between PM levels and health effects being considered (e.g., weather-related factors; behavioral factors such as heating/air conditioning use; driving patterns; and time-activity patterns)?

The Integrated Review Plan for the National Ambient Air Quality Standards for Particulate Matter identifies a series of policy-relevant questions that provide a framework for this assessment of the scientific evidence (U.S. EPA, 2008, 157072). These questions frame the entire review of the NAAQS for PM, and thus are informed by both science and policy considerations. The ISA organizes and presents the scientific evidence such that, when considered along with findings from risk analyses and policy considerations, will help the EPA address these questions during the NAAQS review for PM. In evaluating

In evaluating evidence on welfare effects of PM, the focus will be on evidence that can help inform these questions from the Integrated Review Plan:

What new evidence is available on the relationship between PM mass/size fraction and/or specific PM components and visibility impairment and climate-related and other welfare effects?

To what extent has key scientific evidence now become available to improve our understanding of the nature and magnitude of visibility, climate, and ecosystem responses to PM and the variability associated with those responses (including ecosystem type, climatic conditions, environmental effects and interactions with other environmental factors and pollutants)?

Do the evidence, the air quality assessment, and the risk/exposure assessment provide support for considering alternative averaging times?

At what levels of ambient PM do visibility impairment and/or environmental effects of concern occur? Is there evidence for the occurrence of adverse visibility and other welfare-related effects at levels of PM lower than those observed previously? If so, at what levels and what are the important uncertainties associated with the evidence?

Do the analyses suggest that PM-induced visibility impairment and/or other welfare- effects will occur with current ambient levels of PM or with levels that just meet the current standards? If so, are these effects of sufficient magnitude and/or frequency such that these effects might reasonably be judged to be important from a public welfare perspective? What are the uncertainties associated with these estimates?

What new evidence and/or techniques are available to quantify the benefits of improved visibility and/or other welfare-related effects?

To what extent have important uncertainties identified in the last review been reduced and/or have new uncertainties emerged?

Included below is all of the second chapter of Integrated Science Assessment for Particulate Matter - even if you aren't up to reading through the entire report, skimming through just this overview may change your life forever... I've put some key points in bold:

Integrated Science Assessment for Particulate Matter (Final Report)

EPA has released the final Integrated Science Assessment (ISA) for Particulate Matter (PM). This is EPA’s latest evaluation of the scientific literature on the potential human health and welfare effects associated with ambient exposures to particulate matter (PM). The development of this document is part of the Agency's periodic review of the national ambient air quality standards (NAAQS) for PM. The recently completed PM ISA and supplementary annexes, in conjunction with additional technical and policy assessments developed by EPA’s Office of Air and Radiation, will provide the scientific basis to inform EPA decisions related to the review of the current PM NAAQS.

PM is one of six principal (or criteria) pollutants for which EPA has established NAAQS. Periodically, EPA reviews the scientific basis for these standards by preparing an ISA (formerly called an Air Quality Criteria Document). The ISA and supplementary annexes, in conjunction with additional technical and policy assessments, provide the scientific basis for EPA decisions on the adequacy of the current NAAQS and the appropriateness of possible alternative standards. The Clean Air Scientific Advisory Committee (CASAC), an independent science advisory committee whose existence and whose review and advisory functions are mandated by Section 109 (d) (2) of the Clean Air Act, is charged (among other things) with independent scientific review of EPA's air quality criteria.

The first and second drafts of the PM ISA were released on December 22, 2008 and July 31, 2009, respectively, for independent external peer review and public comment. These drafts were reviewed at public meetings of the CASAC PM Review Panel on April 1-2, 2009 and October 5-6, 2009, respectively. This final PM ISA has benefited from the expert comments received at the CASAC meetings and from public comments, and it has been revised accordingly.

The subsequent chapters of this ISA will present the most policy-relevant information related to this review of the NAAQS for PM. This chapter integrates the key findings from the disciplines evaluated in this current assessment of the PM scientific literature, which includes the atmospheric sciences, ambient air data analyses, exposure assessment, dosimetry, health studies (e.g., toxicological, controlled human exposure, and epidemiologic), and welfare effects. The EPA framework for causal determinations described in Chapter 1 has been applied to the body of scientific evidence in order to collectively examine the health or welfare effects attributed to PM exposure in a two-step process.

As described in Chapter 1, EPA assesses the results of recent relevant publications, building upon evidence available during the previous NAAQS reviews, to draw conclusions on the causal relationships between relevant pollutant exposures and health or environmental effects. This ISA uses a five-level hierarchy that classifies the weight of evidence for causation:

Causal relationship

Likely to be a causal relationship

Suggestive of a causal relationship

Inadequate to infer a causal relationship

Not likely to be a causal relationship

Beyond judgments regarding causality are questions relevant to quantifying health or environmental risks based on our understanding of the quantitative relationships between pollutant exposures and health or welfare effects. Once a determination is made regarding the causal relationship between the pollutant and outcome category, important questions regarding quantitative relationships include:What is the concentration-response or dose-response relationship?Under what exposure conditions (amount deposited, dose or concentration, duration and pattern) are effects observed?What populations appear to be differentially affected (i.e., more susceptible) to effects?What elements of the ecosystem (e.g., types, regions, taxonomic groups, populations, functions, etc.) appear to be affected, or are more sensitive to effects?

To address these questions, in the second step of the EPA framework, the entirety of quantitative evidence is evaluated to identify and characterize potential concentration-response relationships. This requires evaluation of levels of pollutant and exposure durations at which effects were observed for exposed populations including potentially susceptible populations.

This chapter summarizes and integrates the newly available scientific evidence that best informs consideration of the policy-relevant questions that frame this assessment, presented in Chapter 1. Section 2.1 discusses the trends in ambient concentrations and sources of PM and provides a brief summary of ambient air quality. Section 2.2 presents the evidence regarding personal exposure to ambient PM in outdoor and indoor microenvironments, and it discusses the relationship between ambient PM concentrations and exposure to PM from ambient sources. Section 2.3 integrates the evidence for studies that examine the health effects associated with short-and long-term exposure to PM and discusses important uncertainties identified in the interpretation of the scientific evidence. Section 2.4 provides a discussion of policy-relevant considerations, such as potentially susceptible populations, lag structure, and the PM concentration-response relationship, and PM sources and constituents linked to health effects. Section 2.5 summarizes the evidence for welfare effects related to PM exposure. Finally, Section 2.6 provides all of the causal determinations reached for each of the health outcomes and PM exposure durations evaluated in this ISA.

Recently, advances in understanding the spatiotemporal distribution of PM mass and itsconstituents have been made, particularly with regard to PM 2.5 and its components as well asultrafine particles (UFPs). Emphasis in this ISA is placed on the period from 2005-2007,incorporating the most recent validated EPA Air Quality System (AQS) data. The AQS is EPA’srepository for ambient monitoring data reported by the national, and state and local air monitoringnetworks. Measurements of PM 2.5 and PM 10 are reported into AQS, while PM 10-2.5 concentrationsare obtained as the difference between PM 10 and PM 2.5 (after converting PM 10 concentrations fromSTP to local conditions; Section 3.5). Note, however, that a majority of U.S. counties were notrepresented in AQS because their population fell below the regulatory monitoring threshold.Moreover, monitors reporting to AQS were not uniformly distributed across the U.S. or withincounties, and conclusions drawn from AQS data may not apply equally to all parts of a geographicregion. Furthermore, biases can exist for some PM constituents (and hence total mass) owing tovolatilization losses of nitrates and other semi-volatile compounds, and, conversely, to retention ofparticle-bound water by hygroscopic species. The degree of spatial variability in PM was likely to beregion-specific and strongly influenced by local sources and meteorological and topographicconditions.

2.1.1.1. Spatial Variability across the U.S.

AQS data for daily average concentrations of PM 2.5 for 2005-2007 showed considerablevariability across the U.S. (Section 3.5.1.1). Counties with the highest average concentrations ofPM 2.5 (>18 μg/m3) were reported for several counties in the San Joaquin Valley and inland southernCalifornia as well as Jefferson County, AL (containing Birmingham) and Allegheny County, PA(containing Pittsburgh). Relatively few regulatory monitoring sites have the appropriate co-locatedmonitors for computing PM 10-2.5 , resulting in poor geographic coverage on a national scale(Figure 3-10). Although the general understanding of PM differential settling leads to an expectationof greater spatial heterogeneity in the PM 10-2.5 fraction, deposition of particles as a function of sizedepends strongly on local meteorological conditions. Better geographic coverage is available forPM 10 , where the highest reported annual average concentrations (>50 μg/m3) occurred in southernCalifornia, southern Arizona and central New Mexico. The size distribution of PM variedsubstantially by location, with a generally larger fraction of PM 10 mass in the PM 10-2.5 size range inwestern cities (e.g., Phoenix and Denver) and a larger fraction of PM 10 in the PM 2.5 size range ineastern U.S. cities (e.g., Pittsburgh and Philadelphia). UFPs are not measured as part of AQS or anyother routine regulatory network in the U.S. Therefore, limited information is available regardingregional variability in the spatiotemporal distribution of UFPs.

Spatial variability in PM 2.5 components obtained from the Chemical Speciation Network(CSN) varied considerably by species from 2005-2007 (Figures 3-12 through 3-18). The highestannual average organic carbon (OC) concentrations were observed in the western and southeasternU.S. OC concentrations in the western U.S. peaked in the fall and winter, while OC concentrations inthe Southeast peaked anytime between spring and fall. Elemental carbon (EC) exhibited lessseasonality than OC and showed lowest seasonal variability in the eastern half of the U.S. TheDecember 2009 highest annual average EC concentrations were present in Los Angeles, Pittsburgh, New York, andEl Paso. Concentrations of sulfate (SO 4 2–) were higher in the eastern U.S. as a result of higher SO 2emissions in the East compared with the West. There is also considerable seasonal variability withhigher SO 4 2– concentrations in the summer months when the oxidation of SO 2 proceeds at a fasterrate than during the winter. Nitrate (NO 3 –) concentrations were highest in California and during thewinter in the Upper Midwest. In general, NO 3 – was higher in the winter across the country, in part asa result of temperature-driven partitioning and volatilization. Exceptions existed in Los Angeles andRiverside, CA, where high NO 3 – concentrations appeared year-round. There is variation in bothPM 2.5 mass and composition among cities, some of which might be due to regional differences inmeteorology, sources, and topography.

2.1.1.2. Spatial Variability on the Urban and Neighborhood Scales

In general, PM 2.5 has a longer atmospheric lifetime than PM 10-2.5 . As a result, PM 2.5 is morehomogeneously distributed than PM 10-2.5 , whose concentrations more closely reflect proximity tolocal sources (Section 3.5.1.2). Because PM 10 encompasses PM 10-2.5 in addition to PM 2.5 , it alsoexhibits more spatial heterogeneity than PM 2.5 . Urban- and neighborhood-scale variability in PMmass and composition was examined by focusing on 15 metropolitan areas, which were chosenbased on their geographic distribution and coverage in recent health effects studies. The urban areasselected were Atlanta, Birmingham, Boston, Chicago, Denver, Detroit, Houston, Los Angeles, NewYork, Philadelphia, Phoenix, Pittsburgh, Riverside, Seattle and St. Louis. Inter-monitor correlationremained higher over long distances for PM 2.5 as compared with PM 10 in these 15 urban areas. To alarge extent, greater variation in PM 2.5 and PM 10 concentrations within cities was observed in areaswith lower ratios of PM 2.5 to PM 10 . When the data was limited to only sampler pairs with less than4 km separation (i.e., on a neighborhood scale), inter-sampler correlations remained higher for PM 2.5than for PM 10 . The average inter-sampler correlation was 0.93 for PM 2.5 , while it dropped to 0.70 forPM 10 (Section 3.5.1.3). Insufficient data were available in the 15 metropolitan areas to performsimilar analyses for PM 10-2.5 using co-located, low volume FRM monitors.

As previously mentioned, UFPs are not measured as part of AQS or any other routineregulatory network in the U.S. Therefore, information about the spatial variability of UFPs is sparse;however, their number concentrations are expected to be highly spatially and temporally variable.This has been shown on the urban scale in studies in which UFP number concentrations drop offquickly with distance from roads compared to accumulation mode particle numbers.

2.1.2. Trends and Temporal Variability

Overall, PM 2.5 concentrations decreased from 1999 (the beginning of nationwide monitoringfor PM 2.5 ) to 2007 in all ten EPA Regions, with the 3-yr avg of the 98th percentile of 24-h PM 2.5concentrations dropping 10% over this time period. However from 2002-2007, concentrations ofPM 2.5 were nearly constant with decreases observed in only some EPA Regions (Section 3.5.2.1).Concentrations of PM 2.5 components were only available for 2002-2007 using CSN data and showedlittle decline over this time period. This trend in PM 2.5 components is consistent with trends in PM 2.5mass concentration observed after 2002 (shown in Figures 3-44 through 3-47). Concentrations ofPM 10 also declined from 1988 to 2007 in all ten EPA Regions.

Using hourly PM observations in the 15 metropolitan areas, diel variation showed averagehourly peaks that differ by size fraction and region (Section 3.5.2.3). For both PM 2.5 and PM 10 , amorning peak was typically observed starting at approximately 6:00 a.m., corresponding with thestart of morning rush hour. There was also an evening concentration peak that was broader than themorning peak and extended into the overnight period, reflecting the concentration increase caused bythe usual collapse of the mixing layer after sundown. The magnitude and duration of these peaksvaried considerably by metropolitan area investigated.

UFPs were found to exhibit similar two-peaked diel patterns in Los Angeles and the SanJoaquin Valley of CA and Rochester, NY as well as in Kawasaki City, Japan, and Copenhagen,Denmark. The morning peak in UFPs likely represents primary source emissions, such as rush-hourtraffic, while the afternoon peak likely represents the combination of primary source emissions andnucleation of new particles.

2.1.3. Correlations between Copollutants

Correlations between PM and gaseous copollutants, including SO 2 , NO 2 , carbon monoxide(CO) and O 3 , varied both seasonally and spatially between and within metropolitan areas(Section 3.5.3). On average, PM 2.5 and PM 10 were correlated with each other better than with thegaseous copollutants. Although data are limited for PM 10-2.5 , the available data suggest a strongercorrelation between PM 10 and PM 10-2.5 than between PM 2.5 and PM 10-2.5 on a national basis.Therewas relatively little seasonal variability in the mean correlation between PM in both size fractionsand SO 2 and NO 2 . CO, however, showed higher correlations with PM 2.5 and PM 10 on average in thewinter compared with the other seasons. This seasonality results in part because a larger fraction ofPM is primary in origin during the winter. To the extent that this primary component of PM isassociated with common combustion sources of NO 2 and CO, then higher correlations with thesegaseous copollutants are to be expected. Increased atmospheric stability in colder months also resultsin higher correlations between primary pollutants (Section 3.5).

The correlation between daily maximum 8-h avg O 3 and 24-h avg PM 2.5 showed the highestdegree of seasonal variability with positive correlations on average in summer (avg = 0.56) andnegative correlations on average in the winter (avg = -0.30). During the transition seasons, springand fall, correlations were mixed but on average were still positive. PM 2.5 is both primary andsecondary in origin, whereas O 3 is only secondary. Photochemical production of O 3 and secondaryPM in the planetary boundary layer (PBL) is much slower during the winter than during otherseasons. Primary pollutant concentrations (e.g., primary PM 2.5 components, NO and NO 2 ) in manyurban areas are elevated in winter as the result of heating emissions, cold starts and low mixingheights. O 3 in the PBL during winter is mainly associated with air subsiding from above theboundary layer following the passage of cold fronts, and this subsiding air has much lower PMconcentrations than are present in the PBL. Therefore, a negative association between O 3 and PM 2.5is frequently observed in the winter. During summer, both O 3 and secondary PM 2.5 are produced inthe PBL and in the lower free troposphere at faster rates compared to winter, and so they tend to bepositively correlated.

2.1.4. Measurement Techniques

The federal reference methods (FRMs) for PM 2.5 and PM 10 are based on criteria outlined inthe Code of Federal Regulations. They are, however, subject to several limitations that should bekept in mind when using compliance monitoring data for health studies. For example, FRMtechniques are subject to the loss of semi-volatile species such as organic compounds andammonium nitrate (especially in the West). Since FRMs based on gravimetry use 24-h integratedfilter samples to collect PM mass, no information is available for variations over shorter averagingtimes from these instruments. However, methods have been developed to measure real-time PMmass concentrations. Real-time (or continuous and semi-continuous) measurement techniques arealso available for PM species, such as particle into liquid sampler (PILS) for multiple ions analysisand aerosol mass spectrometer (AMS) for multiple components analysis (Section 3.4.1). Advanceshave also been achieved in PM organic speciation. New 24-h FRMs and Federal Equivalent Methods(FEMs) based on gravimetry and continuous FEMs for PM 10-2.5 are available. FRMs for PM 10-2.5 relyon calculating the difference between co-located PM 10 and PM 2.5 measurements while adichotomous sampler is designated as an FEM.

2.1.5. PM Formation in the Atmosphere and Removal

PM in the atmosphere contains both primary (i.e., emitted directly by sources) and secondarycomponents, which can be anthropogenic or natural in origin. Secondary PM components can beproduced by the oxidation of precursor gases such as SO 2 and NO X to acids followed byneutralization with ammonia (NH 3 ) and the partial oxidation of organic compounds. In addition tobeing emitted as primary particles, UFPs are produced by the nucleation of H 2 SO 4 vapor, H 2 Ovapor, and perhaps NH 3 and certain organic compounds. Over most of the earth’s surface, nucleationis probably the major mechanism forming new UFPs. New UFP formation has been observed inenvironments ranging from relatively unpolluted marine and continental environments to pollutedurban areas as an ongoing background process and during nucleation events. However, as notedabove, a large percentage of UFPs come from combustion-related sources such as motor vehicles.

Developments in the chemistry of formation of secondary organic aerosol (SOA) indicate thatoligomers are likely a major component of OC in aerosol samples. Recent observations also suggestthat small but significant quantities of SOA are formed from the oxidation of isoprene in addition tothe oxidation of terpenes and organic hydrocarbons with six or more carbon atoms. Gasoline engineshave been found to emit a mix of nucleation-mode heavy and large polycyclic aromatichydrocarbons on which unspent fuel and trace metals can condense, while diesel particles arecomposed of a soot nucleus on which sulfates and hydrocarbons can condense. To the extent that theprimary component of organic aerosol is overestimated in emissions from combustion sources, thesemi-volatile components are underestimated. This situation results from the lack of capture ofevaporated semi-volatile components upon dilution in common emissions tests. As a result, near-traffic sources of precursors to SOA would be underestimated. The oxidation of these precursorsresults in more oxidized forms of SOA than previously considered, in both near source urbanenvironments and further downwind. Primary organic aerosol can also be further oxidized to formsthat have many characteristics in common with oxidized SOA formed from gaseous precursors.Organic peroxides constitute a significant fraction of SOA and represent an important class ofreactive oxygen species (ROS) that have high oxidizing potential. More information on sources,emissions and deposition of PM are included in Section 3.3.

Wet and dry deposition are important processes for removing PM and other pollutants from theatmosphere on urban, regional, and global scales. Wet deposition includes incorporation of particlesinto cloud droplets that fall as rain (rainout) and collisions with falling rain (washout). Otherhydrometeors (snow, ice) can also serve the same purpose. Dry deposition involves transfer ofparticles through gravitational settling and/or by impaction on surfaces by turbulent motions. Theeffects of deposition of PM on ecosystems and materials are discussed in Section 2.5 and inChapter 9.

2.1.6. Source Contributions to PM

Results of receptor modeling calculations indicate that PM 2.5 is produced mainly bycombustion of fossil fuel, either by stationary sources or by transportation. A relatively small numberof broadly defined source categories, compared to the total number of chemical species that typicallyare measured in ambient monitoring source receptor studies, account for the majority of the observedPM mass. Some ambiguity is inherent in identifying source categories. For example, quite differentmobile sources such as trucks, farm equipment, and locomotives rely on diesel engines and ancillarydata is often required to resolve these sources. A compilation of study results shows that secondarySO 4 2– (derived mainly from SO 2 emitted by Electricity Generating Units [EGUs]), NO 3 – (from theoxidation of NO x emitted mainly from transportation sources and EGUs), and primary mobile sourcecategories, constitute most of PM 2.5 (and PM 10 ) in the East. PM 10-2.5 is mainly primary in origin,having been emitted as fully formed particles derived from abrasion and crushing processes, soildisturbances, plant and insect fragments, pollens and other microorganisms, desiccation of marineaerosol emitted from bursting bubbles, and hygroscopic fine PM expanding with humidity to coarsemode. Gases such as HNO 3 can also condense directly onto preexisting coarse particles. Suspendedprimary coarse PM can contain Fe, Si, Al, and base cations from soil, plant and insect fragments,pollen, fungal spores, bacteria, and viruses, as well as fly ash, brake lining particles, debris, andautomobile tire fragments. Quoted uncertainties in the source apportionment of constituents inambient aerosol samples typically range from 10 to 50%. An intercomparison of sourceapportionment techniques indicated that the same major source categories of PM 2.5 were consistentlyidentified by several independent groups working with the same data sets. Soil-, sulfate-, residualoil-, and salt-associated mass were most clearly identified by the groups. Other sources with moreambiguous signatures, such as vegetative burning and traffic-related emissions were less consistentlyidentified.

Spatial variability in source contributions across urban areas is an important consideration inassessing the likelihood of exposure error in epidemiologic studies relating health outcomes tosources. Concepts similar to those for using ambient concentrations as surrogates for personalexposures apply here. Some source attribution studies for PM 2.5 indicate that intra-urban variabilityincreases in the following order: regional sources (e.g., secondary SO 4 2– originating from EGUs)< area sources (e.g., on-road mobile sources) < point sources (e.g., metals from stacks of smelters).

Although limited information was available for PM 10-2.5 , it does indicate a similar ordering, butwithout a regional component (resulting from the short lifetime of PM 10-2.5 compared to transporttimes on the regional scale). More discussion on source contributions to PM is available inSection 3.6.

2.1.7. Policy-Relevant Background

The background concentrations of PM that are useful for risk and policy assessments, whichinform decisions about the NAAQS are referred to as policy-relevant background (PRB)concentrations. PRB concentrations have historically been defined by EPA as those concentrationsthat would occur in the U.S. in the absence of anthropogenic emissions in continental North Americadefined here as the U.S., Canada, and Mexico. For this document, PRB concentrations includecontributions from natural sources everywhere in the world and from anthropogenic sources outsidecontinental North America. Background concentrations so defined facilitated separation of pollutionthat can be controlled by U.S. regulations or through international agreements with neighboringcountries from those that were judged to be generally uncontrollable by the U.S. Over time,consideration of potential broader ranging international agreements may lead to alternativedeterminations of which PM source contributions should be considered by EPA as part of PRB.

Contributions to PRB concentrations of PM include both primary and secondary natural andanthropogenic components. For this document, PRB concentrations of PM 2.5 for the continental U.S.were estimated using EPA’s Community Multi-scale Air Quality (CMAQ) modeling system, adeterministic, chemical-transport model (CTM), using output from GEOS-Chem a global-scalemodel for CMAQ boundary conditions. PRB concentrations of PM 2.5 were estimated to be less than1 μg/m3 on an annual basis, with maximum daily average values in a range from 3.1 to 20 μg/m3 andhaving a peak of 63 μg/m3 at the nine national park sites across the U.S. used to evaluate modelperformance for this analysis. A description of the models and evaluation of their performance isgiven in Section 3.6 and further details about the calculations of PRB concentrations are given inSection 3.7.

2.2. Human Exposure

This section summarizes the findings from the recent exposure assessment literature. Thissummary is intended to support the interpretation of the findings from epidemiologic studies andreflects the material presented in Section 3.8. Attention is given to how concentration metrics can beused in exposure assessment and what errors and uncertainties are incurred for different approaches.Understanding of exposure errors is important because exposure error can potentially bias anestimate of a health effect or increase the size of confidence intervals around a health effect estimate.2.2.1. Spatial Scales of PM Exposure Assessment

Assessing population-level exposure at the urban scale is particularly relevant for time-seriesepidemiologic studies, which provide information on the relationship between health effects andcommunity-average exposure, rather than an individual’s exposure. PM concentrations measured at acentral-site ambient monitor are used as surrogates for personal PM exposure. However, thecorrelation between the PM concentration measured at central-site ambient monitor(s) and theunknown true community average concentration depends on the spatial distribution of PM, thelocation of the monitoring site(s) chosen to represent the community average, and division of thecommunity by terrain features or local sources into several sub-communities that differ in thetemporal pattern of pollution. Concentrations of SO 4 2– and some components of SOA measured atcentral-site monitors are expected to be uniform in urban areas because of the regional nature of theirsources. However, this is not true for primary components like EC whose sources are stronglyspatially variable in urban areas.

At micro-to-neighborhood scales, heterogeneity of sources and topography contribute tovariability in exposure. This is particularly true for PM 10-2.5 and for UFPs, which have spatiallyvariable urban sources and loss processes (mainly gravitational settling for PM 10-2.5 and coagulationfor UFPs) that also limit their transport from sources more readily than for PM 2.5 . Personal activitypatterns also vary across urban areas and across regions. Some studies, conducted mainly in Europe,have found personal PM 2.5 and PM 10 exposures for pedestrians in street canyons to be higher thanambient concentrations measured by urban central site ambient monitors. Likewise,microenvironmental UFP concentrations were observed to be substantially higher in near-roadenvironments, street canyons, and tunnels when compared with urban background concentrations.In-vehicle UFP and PM 2.5 exposures can also be important. As a result, concentrations measured byambient monitors likely do not reflect the contributions of UFP or PM 2.5 exposures to individualswhile commuting.

There is significant variability within and across regions of the country with respect to indoorexposures to ambient PM. Infiltrated ambient PM concentrations depend in part on the ventilationproperties of the building or vehicle in which the person is exposed. PM infiltration factors dependon particle size, chemical composition, season, and region of the country. Infiltration can best bemodeled dynamically rather than being represented by a single value. Season is important to PMinfiltration because it affects the ventilation practices (e.g., open windows) used. In addition, ambienttemperature and humidity conditions affect the transport, dispersion, and size distribution of PM.Residential air exchange rates have been observed to be higher in the summer for regions with lowair conditioning usage. Regional differences in air exchange rates (Southwest < Southeast< Northeast < Northwest) also reflect ventilation practices. Differential infiltration occurs as afunction of PM size and composition (the latter of which is described below). PM infiltration islarger for accumulation mode particles than for UFPs and PM 10-2.5 . Differential infiltration by sizefraction can affect exposure estimates if not accurately characterized.

2.2.2. Exposure to PM Components and Copollutants

Emission inventories and source apportionment studies suggest that sources of PM exposurevary by region. Comparison of studies performed in the eastern U.S. with studies performed in thewestern U.S. suggest that the contribution of SO 4 2– to exposure is higher for the East (16-46%)compared with the West (~4%) and that motor vehicle emissions and secondary NO 3 – are largersources of exposure for the West (~9%) as compared with the East (~4%). Results of sourceapportionment studies of exposure to SO 4 2– indicate that SO 4 2– exposures are mainly attributable toambient sources. Source apportionment for OC and EC is difficult because they originate from bothindoor and outdoor sources. Exposure to OC of indoor and outdoor origin can be distinguished bythe presence of aliphatic C-H groups generated indoors, since outdoor concentrations of aliphaticC-H are low. Studies of personal exposure to ambient trace metal have shown significant variationamong cities and over seasons. This is in response to geographic and seasonal variability in sourcesincluding incinerator operation, fossil fuel combustion, biomass combustion (wildfires), and theresuspension of crustal materials in the built environment. Differential infiltration is also affected byvariations in particle composition and volatility. For example, EC infiltrates more readily than OC.This can lead to outdoor-indoor differentials in PM composition.

Some studies have explored the relationship between PM and copollutant gases and suggestedthat certain gases can serve as surrogates for describing exposure to other air pollutants. The findingsindicate that ambient concentrations of gaseous copollutants can act as surrogates for personalexposure to ambient PM. Several studies have concluded that ambient concentrations of O 3 , NO 2 ,and SO 2 are associated with the ambient component of personal exposure to total PM 2.5 . Ifassociations between ambient gases and personal exposure to PM 2.5 of ambient origin exist, suchassociations are complex and vary by season and location.2.2.3. Implications for Epidemiologic Studies

In epidemiologic studies, exposure may be estimated using various approaches, most of whichrely on measurements obtained using central site monitors. The magnitude and direction of thebiases introduced through error in exposure measurement depend on the extent to which the error isassociated with the measured PM concentration. In general, when exposure error is not stronglycorrelated with the measured PM concentration, bias is toward the null and effect estimates areunderestimated. Moreover, lack of information regarding exposure measurement error can also adduncertainty to the health effects estimate.

One important factor to be considered is the spatial variation in PM concentrations. The degreeof urban-scale spatial variability in PM concentrations varies across the country and by size fraction.PM 2.5 concentrations are relatively well-correlated across monitors in the urban areas examined forthis assessment. The limited available evidence indicates that there is greater spatial variability inPM 10-2.5 concentrations than PM 2.5 concentrations, resulting in increased exposure error for the largersize fraction. Likewise, studies have shown UFPs to be more spatially variable across urban areascompared to PM 2.5 . Even if PM 2.5 , PM 10-2.5 , or UFP concentrations measured at sites within an urbanarea are generally highly correlated, significant spatial variation in their concentrations can occur onany given day. In addition, there can be differential exposure errors for PM components (e.g., SO 4 2–,OC, EC). Current information suggests that UFPs, PM 10-2.5, and some PM components are morespatially variable than PM 2.5 . Spatial variability of these PM indicators adds uncertainty to exposureestimates.

Overall, recent studies generally confirm and build upon the key conclusions of the 2004 PMAQCD: separation of total PM exposures into ambient and nonambient components reducespotential uncertainties in the analysis and interpretation of PM health effects data; and ambient PMconcentration can be used as a surrogate for ambient PM exposure in community time-seriesepidemiologic studies because the change in ambient PM concentration should be reflected in thechange in the health risk coefficient. The use of the community average ambient PM 2.5 concentrationas a surrogate for the community average personal exposure to ambient PM 2.5 is not expected tochange the principal conclusions from time-series and most panel epidemiologic studies that usecommunity average health and pollution data. Several recent studies support this by showing howthe ambient component of personal exposure to PM 2.5 could be estimated using various tracer andsource apportionment techniques and by showing that the ambient component is highly correlatedwith ambient concentrations of PM 2.5 . These studies show that the non-ambient component ofpersonal exposure to PM 2.5 is largely uncorrelated with ambient PM 2.5 concentrations. A few panelepidemiologic studies have included personal as well as ambient monitoring data, and generallyreported associations with all types of PM measurements. Epidemiologic studies of long-termexposure typically exploit the differences in PM concentration across space, as well as time, toestimate the effect of PM on the health outcome of interest. Long-term exposure estimates are mostaccurate for pollutants that do not vary substantially within the geographic area studied.

2.3. Health Effects

This section evaluates the evidence from toxicological, controlled human exposure, andepidemiologic studies that examined the health effects associated with short- and long-term exposureto PM (i.e., PM 2.5 , PM 10-2.5 and UFPs). The results from the health studies evaluated in combinationwith the evidence from atmospheric chemistry and exposure assessment studies contribute to thecausal determinations made for the health outcomes discussed in this assessment (a description ofthe causal framework can be found in Section 1.5.4). In the following sections a discussion of thecausal determinations will be presented by PM size fraction and exposure duration (i.e., short- orlong-term exposure) for the health effects for which sufficient evidence was available to conclude acausal, likely to be causal or suggestive relationship. Although not presented in depth in this chapter,a detailed discussion of the underlying evidence used to formulate each causal determination can befound in Chapters 6 and 7.

Controlled human exposure studies have demonstrated PM 2.5 -induced changes in variousmeasures of cardiovascular function among healthy and health-compromised adults. The mostconsistent evidence is for altered vasomotor function following exposure to diesel exhaust (DE) orCAPs with O 3 (Section 6.2.4.2). Although these findings provide biological plausibility for theobservations from epidemiologic studies, the fresh DE used in the controlled human exposurestudies evaluated contains gaseous components (e.g., CO, NO x ), and therefore, the possibility thatsome of the changes in vasomotor function might be due to gaseous components cannot be ruled out.Furthermore, the prevalence of UFPs in fresh DE limits the ability to conclusively attribute theobserved effects to either the UF fraction or PM 2.5 as a whole. An evaluation of toxicological studiesfound evidence for altered vessel tone and microvascular reactivity, which provide coherence andbiological plausibility for the vasomotor effects that have been observed in both the controlledhuman exposure and epidemiologic studies (Section 6.2.4.3). However, most of these toxicologicalstudies exposed animals via intratracheal (IT) instillation or using relatively high inhalationconcentrations.

Controlled human exposure studies using adult volunteers have demonstrated increasedmarkers of pulmonary inflammation following exposure to a variety of different particle types;oxidative responses to DE and wood smoke; and exacerbations of allergic responses and allergicsensitization following exposure to DE particles (Section 6.3). Toxicological studies have providedadditional support for PM 2.5 -related respiratory effects through inhalation exposures of animals toCAPs, DE, other traffic-related PM and wood smoke. These studies reported an array of respiratoryeffects including altered pulmonary function, mild pulmonary inflammation and injury, oxidativeresponses, airway hyperresponsiveness (AHR) in allergic and non-allergic animals, exacerbations ofallergic responses, and increased susceptibility to infections (Section 6.3).

Overall, the evidence for an effect of PM 2.5 on respiratory outcomes is somewhat restricted bylimited coherence between some of the findings from epidemiologic and controlled human exposurestudies for the specific health outcomes reported and the sub-populations in which those healthoutcomes occur. Epidemiologic studies have reported variable results among specific respiratoryoutcomes, specifically in asthmatics (e.g., increased respiratory symptoms in asthmatic children, butnot increased asthma hospital admissions and ED visits) (Section 6.3.8). Additionally, respiratoryeffects have not been consistently demonstrated following controlled exposures to PM 2.5 amongasthmatics or individuals with COPD. Collectively, the epidemiologic, controlled human exposure,and toxicological studies evaluated demonstrate a wide range of respiratory responses, and althoughresults are not fully consistent and coherent across studies the evidence is sufficient to conclude thata causal relationship is likely to exist between short-term exposures to PM 2.5 andrespiratory effects.

Mortality

An evaluation of the epidemiologic literature indicates consistent positive associationsbetween short-term exposure to PM 2.5 and all-cause, cardiovascular-, and respiratory-relatedmortality (Section 6.5.2.2.). The evaluation of multicity studies found that consistent and precise riskestimates for all-cause (nonaccidental) mortality that ranged from 0.29 to 1.21% per 10 μg/m3increase in PM 2.5 at lags of 1 and 0-1 days. In these study locations, mean 24-h avg PM 2.5concentrations were 12.8 μg/m3 and above (Table 6-15). Cardiovascular-related mortality riskestimates were found to be similar to those for all-cause mortality; whereas, the risk estimates forrespiratory-related mortality were consistently larger (i.e., 1.01-2.2%) using the same lag periods andaveraging indices. The studies evaluated that examined the relationship between short-term exposureto PM 2.5 and cardiovascular effects (Section 6.2) provide coherence and biological plausibility forPM 2.5 -induced cardiovascular mortality, which represents the largest component of total(nonaccidental) mortality (~ 35%) (American Heart Association, 2009, 198920). However, as notedin Section 6.3, there is limited coherence between some of the respiratory morbidity findings fromepidemiologic and controlled human exposure studies for the specific health outcomes reported andthe subpopultions in which those health outcomes occur, complicating the interpretation of the PM 2.5respiratory mortality effects observed. Regional and seasonal patterns in PM 2.5 risk estimates wereobserved with the greatest effect estimates occurring in the eastern U.S. and during the spring. Of thestudies evaluated only Burnett et al. (2004, 086247), a Canadian multicity study, analyzed gaseouspollutants and found mixed results, with possible confounding of PM 2.5 risk estimates by NO 2 .Although the recently evaluated U.S.-based multicity studies did not analyze potential confoundingof PM 2.5 risk estimates by gaseous pollutants, evidence from the limited number of single-citystudies evaluated in the 2004 PM AQCD (U.S. EPA, 2004, 056905) suggest that gaseouscopollutants do not confound the PM 2.5 -mortality association. This is further supported by studiesthat examined the PM 10 -mortality relationship. An examination of effect modifiers (e.g.,demographic and socioeconomic factors), specifically air conditioning use as an indicator fordecreased pollutant penetration indoors, has suggested that PM 2.5 risk estimates increase as thepercent of the population with access to air conditioning decreases. Collectively, the epidemiologicliterature provides evidence that a causal relationship exists between short-term exposuresto PM 2.5 and mortality.

2.3.1.2. Effects of Long-Term Exposure to PM 2.5

Cardiovascular Effects

The strongest evidence for cardiovascular health effects related to long-term exposure to PM 2.5comes from large, multicity U.S.-based studies, which provide consistent evidence of an associationbetween long-term exposure to PM 2.5 and cardiovascular mortality (Section 7.2.10). Theseassociations are supported by a large U.S.-based epidemiologic study (i.e., Women’s Health Initiative[WHI] study) that reports associations between PM 2.5 and CVDs among post-menopausal womenusing a 1-yr avg PM 2.5 concentration (mean = 13.5 μg/m3) (Section 7.2). However, epidemiologicstudies that examined subclinical markers of CVD report inconsistent findings. Epidemiologicstudies have also provided some evidence for potential modification of the PM 2.5 -CVD associationwhen examining individual-level data, specifically smoking status and the use of anti-hyperlipidemics. Although epidemiologic studies have not consistently detected effects on markersof atherosclerosis due to long-term exposure to PM 2.5 , toxicological studies have provided strongevidence for accelerated development of atherosclerosis in ApoE-/- mice exposed to CAPs and haveshown effects on coagulation, experimentally-induced hypertension, and vascular reactivity (Section7.2.1.2). Evidence from toxicological studies provides biological plausibility and coherence withstudies of short-term exposure and cardiovascular morbidity and mortality, as well as with studiesthat examined long-term exposure to PM 2.5 and cardiovascular mortality. Taken together, theevidence from epidemiologic and toxicological studies is sufficient to conclude that a causalrelationship exists between long-term exposures to PM 2.5 and cardiovascular effects.

Respiratory Effects

Recent epidemiologic studies conducted in the U.S. and abroad provide evidence ofassociations between long-term exposure to PM 2.5 and decrements in lung function growth,increased respiratory symptoms, and asthma development in study locations with mean PM 2.5concentrations ranging from 13.8 to 30 μg/m3 during the study periods (Section 7.3.1.1 and Section7.3.2.1). These results are supported by studies that observed associations between long-termexposure to PM 10 and an increase in respiratory symptoms and reductions in lung function growth inareas where PM 10 is dominated by PM 2.5 . However, the evidence to support an association withlong-term exposure to PM 2.5 and respiratory mortality is limited (Figure 7-7). Subchronic andchronic toxicological studies of CAPs, DE, roadway air and woodsmoke provide coherence andbiological plausibility for the effects observed in the epidemiologic studies. These toxicologicalstudies have presented some evidence for altered pulmonary function, mild inflammation, oxidativeresponses, immune suppression, and histopathological changes including mucus cell hyperplasia(Section 7.3). Exacerbated allergic responses have been demonstrated in animals exposed to DE andwood smoke. In addition, pre- and postnatal exposure to ambient levels of urban particles was foundto affect lung development in an animal model. This finding is important because impaired lungdevelopment is one mechanism by which PM exposure may decrease lung function growth inchildren. Collectively, the evidence from epidemiologic and toxicological studies is sufficient toconclude that a causal relationship is likely to exist between long-term exposures to PM 2.5and respiratory effects.

Mortality

The recent epidemiologic literature reports associations between long-term PM 2.5 exposureand increased risk of mortality. Mean PM 2.5 concentrations ranged from 13.2 to 29 μg/m3 during thestudy period in these areas (Section 7.6). When evaluating cause-specific mortality, the strongestevidence can be found when examining associations between PM 2.5 and cardiovascular mortality,and positive associations were also reported between PM 2.5 and lung cancer mortality (Figure 7-7).The cardiovascular mortality association has been confirmed further by the extended Harvard SixCities and American Cancer Society studies, which both report strong associations between long-term exposure to PM 2.5 and cardiopulmonary and IHD mortality (Figure 7-7). Additional newevidence from a study that used the WHI cohort found a particularly strong association betweenlong-term exposure to PM 2.5 and CVD mortality in post-menopausal women. Fewer studies haveevaluated the respiratory component of cardiopulmonary mortality, and, as a result, the evidence tosupport an association with long-term exposure to PM 2.5 and respiratory mortality is limited (Figure7-7). The evidence for cardiovascular and respiratory morbidity due to short- and long-term exposureto PM 2.5 provides biological plausibility for cardiovascular- and respiratory-related mortality.Collectively, the evidence is sufficient to conclude that a causal relationship exists betweenlong-term exposures to PM 2.5 and mortality.

Reproductive and Developmental Effects

Evidence is accumulating for PM 2.5 effects on low birth weight and infant mortality, especiallydue to respiratory causes during the post-neonatal period. The mean PM 2.5 concentrations during thestudy periods ranged from 5.3-27.4 μg/m3 (Section 7.4), with effects becoming more precise andconsistently positive in locations with mean PM 2.5 concentrations of 15 μg/m3 and above(Section 7.4). Exposure to PM 2.5 was usually associated with greater reductions in birth weight thanexposure to PM 10 . The evidence from a few U.S. studies that investigated PM 10 effects on fetalgrowth, which reported similar decrements in birth weight, provide consistency for the PM 2.5associations observed and strengthen the interpretation that particle exposure may be causally relatedto reductions in birth weight. The epidemiologic literature does not consistently report associationsbetween long-term exposure to PM and preterm birth, growth restriction, birth defects or decreasedsperm quality. Toxicological evidence supports an association between PM 2.5 and PM 10 exposure andadverse reproductive and developmental outcomes, but provide little mechanistic information orbiological plausibility for an association between long-term PM exposure and adverse birthoutcomes (e.g., low birth weight or infant mortality). New evidence from animal toxicologicalstudies on heritable mutations is of great interest, and warrants further investigation. Overall, theepidemiologic and toxicological evidence is suggestive of a causal relationship between long-term exposures to PM 2.5 and reproductive and developmental outcomes.

Cancer, Mutagenicity, and Genotoxicity

Multiple epidemiologic studies have shown a consistent positive association between PM 2.5and lung cancer mortality, but studies have generally not reported associations between PM 2.5 andlung cancer incidence (Section 7.5). Animal toxicological studies have examined the potentialrelationship between PM and cancer, but have not focused on specific size fractions of PM. Insteadthey have examined ambient PM, wood smoke, and DEP. A number of studies indicate that ambienturban PM, emissions from wood/biomass burning, emissions from coal combustion, and gasolineand DE are mutagenic, and that PAHs are genotoxic. These findings are consistent with earlierstudies that concluded that ambient PM and PM from specific combustion sources are mutagenic andgenotoxic and provide biological plausibility for the results observed in the epidemiologic studies. Alimited number of epidemiologic and toxicological studies examined epigenetic effects, anddemonstrate that PM induces some changes in methylation. However, it has yet to be determinedhow these alterations in the genome could influence the initiation and promotion of cancer.Additionally, inflammation and immune suppression induced by exposure to PM may confersusceptibility to cancer. Collectively, the evidence from epidemiologic studies, primarily those oflung cancer mortality, along with the toxicological studies that show some evidence of the mutagenicand genotoxic effects of PM is suggestive of a causal relationship between long-termexposures to PM 2.5 and cancer.

2.3.2. Integration of PM 2.5 Health Effects

In epidemiologic studies, short-term exposure to PM 2.5 is associated with a broad range ofrespiratory and cardiovascular effects, as well as mortality. For cardiovascular effects and mortality,the evidence supports the existence of a causal relationship with short-term PM 2.5 exposure; whilethe evidence indicates that a causal relationship is likely to exist between short-term PM 2.5 exposureand respiratory effects. The effect estimates from recent and older U.S. and Canadian-basedepidemiologic studies that examined the relationship between short-term exposure to PM 2.5 andhealth outcomes with mean 24-h avg PM 2.5 concentrations <17 μg/m3 are shown in Figure 2-1. Anumber of different health effects are included in Figure 2-1 to provide an integration of the range ofeffects by mean concentration, with a focus on cardiovascular and respiratory effects and all-cause(nonaccidental) mortality (i.e., health effects categories with at least a suggestive causaldetermination). A pattern of consistent positive associations with mortality and morbidity effects canbe seen in this figure. Mean PM 2.5 concentrations ranged from 6.1 to 16.8 μg/m3.in these studylocations.

Figure 2-1: Summary of effect estimates (per 10 μg/m3) by increasing concentration from U.S. studies examining the association between short-term exposure to PM 2.5 and cardiovascular and respiratory effects, and mortality, conducted in locations where the reported mean 24-h avg PM 2.5 concentrations were <17 μg/m3.

Long-term exposure to PM 2.5 has been associated with health outcomes similar to those foundin the short-term exposure studies, specifically for respiratory and cardiovascular effects andmortality. As found for short-term PM 2.5 exposure, the evidence indicates that a causal relationshipexists between long-term PM 2.5 exposure and cardiovascular effects and mortality, and that a causalrelationship is likely to exist between long-term PM 2.5 exposure and effects on the respiratorysystem.

Figure 2-2 highlights the findings of epidemiologic studies where the long-term mean PM 2.5concentrations were ≤ 29 μg/m3. A range of health outcomes are displayed (including cardiovascularmortality, all-cause mortality, infant mortaltiy, and bronchitis) ordered by mean concentration. Therange of mean PM 2.5 concentrations in these studies was 10.7-29 μg/m3 during the study periods.Additional studies not included in this figure that focus on subclinical outcomes, such as changes inlung function or atherosclerotic markers also report effects in areas with similar concentrations(Sections 7.2 and 7.3). Although not highlighted in the summary figure, long-term PM 2.5 exposurestudies also provide evidence for reproductive and developmental effects (i.e., low birth weight) andcancer (i.e., lung cancer mortality) in response to to exposure to PM 2.5.

Figure 2-2: Summary of effect estimates (per 10 μg/m3) by increasing concentration from U.S. studies examining the association between long-term exposure to PM 2.5 and cardiovascular and respiratory effects, and mortality.

The observations from both the short- and long-term exposure studies are supported byexperimental findings of PM 2.5 -induced subclinical and clinical cardiovascular effects.Epidemiologic studies have shown an increase in ED visits and hospital admissions for IHD uponexposure to PM 2.5 . These effects are coherent with the changes in vasomotor function and ST-segment depression observed in both toxicological and controlled human exposure studies. It hasbeen postulated that exposure to PM 2.5 can lead to myocardial ischemia through an effect on theautonomic nervous system or by altering vasomotor function. PM-induced systemic inflammation,oxidative stress and/or endothelial dysfunction may contribute to altered vasomotor function. Theseeffects have been demonstrated in recent animal toxicological studies, along with alteredmicrovascular reactivity, altered vessel tone, and reduced blood flow during ischemia. Toxicologicalstudies demonstrating increased right ventricular pressure and diminished cardiac contractility alsoprovide biological plausibility for the associations observed between PM 2.5 and CHF inepidemiologic studies.

Thus, the overall evidence from the short-term epidemiologic, controlled human exposure, andtoxicological studies evaluated provide coherence and biological plausibility for cardiovasculareffects related to myocardial ischemia and CHF. Coherence in the cardiovascular effects observedcan be found in long-term exposure studies, especially for CVDs among post-menopausal women.Additional studies provide limited evidence for subclinical measures of atherosclerosis inepidemiologic studies with stronger evidence from toxicological studies that have demonstratedaccelerated development of atherosclerosis in ApoE-/- mice exposed to PM 2.5 CAPs along witheffects on coagulation, experimentally-induced hypertension, and vascular reactivity. Repeated acuteresponses to PM may lead to cumulative effects that manifest as chronic disease, such asatherosclerosis. Contributing factors to atherosclerosis development include systemic inflammation,endothelial dysfunction, and oxidative stress all of which are associated with PM 2.5 exposure.However, it has not yet been determined whether PM initiates or promotes atherosclerosis. Theevidence from both short- and long-term exposure studies on cardiovascular morbidity providecoherence and biological plausibility for the cardiovascular mortality effects observed whenexamining both exposure durations. In addition, cardiovascular hospital admission and mortalitystudies that examined the PM 10 concentration-response relationship found evidence of a log-linearno-threshold relationship between PM exposure and cardiovascular-related morbidity (Section 6.2)and mortality (Section 6.5).

Epidemiologic studies have also reported respiratory effects related to short-term exposure toPM 2.5 , which include increased ED visits and hospital admissions, as well as alterations in lungfunction and respiratory symptoms in asthmatic children. These respiratory effects were found to begenerally robust to the inclusion of gaseous pollutants in copollutant models with the strongestevidence from the higher powered studies (Figure 6-9 and Figure 6-15). Consistent positiveassociations were also reported between short-term exposure to PM 2.5 and respiratory mortality inepidemiologic studies. However, uncertainties exist in the PM 2.5 -respiratory mortality associationsreported due to the limited number of studies that examined potential confounders of the PM 2.5 -respiratory mortality relationship, and the limited information regarding the biological plausibility ofthe clinical and subclinical respiratory outcomes observed in the epidemiologic and controlledhuman exposure studies (Section 6.3) resulting in the progression to PM 2.5 -induced respiratorymortality. Important new findings, which support the PM 2.5 -induced respiratory effects mentionedabove, include associations with post-neonatal (between 1 mo and 1 yr of age) respiratory mortality.Controlled human exposure studies provide some support for the respiratory findings fromepidemiologic studies, with demonstrated increases in pulmonary inflammation following short-termexposure. However, there is limited and inconsistent evidence of effects in response to controlledexposures to PM 2.5 on respiratory symptoms or pulmonary function among healthy adults or adultswith respiratory disease. Long-term exposure epidemiologic studies provide additional evidence forPM 2.5 -induced respiratory morbidity, but little evidence for an association with respiratory mortality.These epidemiologic morbidity studies have found decrements in lung function growth, as well asincreased respiratory symptoms, and asthma. Toxicological studies provide coherence and biologicalplausibility for the respiratory effects observed in response to short and long-term exposures to PMby demonstrating a wide array of biological responses including: altered pulmonary function, mildpulmonary inflammation and injury, oxidative responses, and histopathological changes in animalsexposed by inhalation to PM 2.5 derived from a wide variety of sources. In some cases, prolongedexposures led to adaptive responses. Important evidence was also found in an animal model foraltered lung development following pre- and post-natal exposure to urban air, which may provide amechanism to explain the reduction in lung function growth observed in children in response tolong-term exposure to PM.

Additional respiratory-related effects have been tied to allergic responses. Epidemiologicstudies have provided evidence for increased hospital admissions for allergic symptoms (e.g.,allergic rhinitis) in response to short- and long-term exposure to PM 2.5 . Panel studies also positivelyassociate long-term exposure to PM 2.5 and PM 10 with indicators of allergic sensitization. Controlledhuman exposure and toxicological studies provide coherence for the exacerbation of allergicsymptoms, by showing that PM 2.5 can promote allergic responses and intensify existing allergies.Allergic responses require repeated exposures to antigen over time and co-exposure to an adjuvant(possibly DE particles or UF CAPs) can enhance this response. Allergic sensitization often underliesallergic asthma, characterized by inflammation and AHR. In this way, repeated or chronic exposuresinvolving multifactorial responses (immune system activation, oxidative stress, inflammation) canlead to irreversible outcomes. Epidemiologic studies have also reported evidence for increasedhospital admissions for respiratory infections in response to both short- and long-term exposures toPM 2.5 . Toxicological studies suggest that PM impairs innate immunity, which is the first line ofdefense against infection, providing coherence for the respiratory infection effects observed inepidemiologic studies.

The difference in effects observed across studies and between cities may be attributed, at leastin part, to the differences in PM composition across the U.S. Differences in PM toxicity may resultfrom regionally varying PM composition and size distribution, which in turn reflects differences insources and PM volatility. A person’s exposure to ambient PM will also vary due to regionaldifferences in personal activity patterns, microenvironmental characteristics and the spatialvariability of PM concentrations in urban areas. Regional differences in PM 2.5 composition areoutlined briefly in Section 2.1 above and in more detail in Section 3.5. An examination of data fromthe CSN indicates that East-West gradients exist for a number of PM components. Specifically, SO 4 2-concentrations are higher in the East, OC constitutes a larger fraction of PM in the West, and NO 3 -concentrations are highest in the valleys of central California and during the winter in the Midwest.However, the available evidence and the limited amount of city-specific speciated PM 2.5 data doesnot allow conclusions to be drawn that specifically differentiate effects of PM in different locations.

It remains a challenge to determine relationships between specific constituents, combinationsof constituents, or sources of PM 2.5 and the various health effects observed. Source apportionmentstudies of PM 2.5 have attempted to decipher some of these relationships and in the process haveidentified associations between multiple sources and various respiratory and cardiovascular healtheffects, as well as mortality. Although different source apportionment methods have been used acrossthese studies, the methods used have been evaluated and found generally to identify the samesources and associations between sources and health effects (Section 6.6). While uncertaintyremains, it has been recognized that many sources and components of PM 2.5 contribute to healtheffects. Overall, the results displayed in Table 6-18 indicate that many constituents of PM 2.5 can belinked with multiple health effects, and the evidence is not yet sufficient to allow differentiation ofthose constituents or sources that are more closely related to specific health outcomes.

Variability in the associations observed across PM 2.5 epidemiologic studies may be due in partto exposure error related to the use of county-level air quality data. Because western U.S. countiestend to be much larger and more topographically diverse than eastern U.S. counties, the day-to-dayvariations in concentration at one site, or even for the average of several sites, may not correlate wellwith the day-to-day variations in all parts of the county. For example, site-to-site correlations as afunction of distance between sites (Section 3.5.1.2) fall off rapidly with distance in Los Angeles, buthigh correlations extend to larger distances in eastern cities such as Boston and Pittsburgh. Thesedifferences may be attributed to a number of factors including topography, the built environment,climate, source characteristics, ventilation usage, and personal activity patterns. For instance,regional differences in climate and infrastructure can affect time spent outdoors or indoors, airconditioning usage, and personal activity patterns. Characteristics of housing stock may also causeregional differences in effect estimates because new homes tend to have lower infiltration factorsthan older homes. Biases and uncertainties in exposure estimates resulting from these aspects can, inturn, cause bias and uncertainty in associated health effects estimates.

The new evidence reviewed in this ISA greatly expands upon the evidence available in the2004 PM AQCD particularly in providing greater understanding of the underlying mechanisms forPM 2.5 induced cardiovascular and respiratory effects for both short- and long-term exposures. Recentstudies have provided new evidence linking long-term exposure to PM 2.5 with cardiovascularoutcomes that has expanded upon the continuum of effects ranging from the more subtle subclinicalmeasures to cardiopulmonary mortality.

2.3.3. Exposure to PM10-2.5

2.3.3.1. Effects of Short-Term Exposure to PM 10-2.5

Cardiovascular Effects

Generally positive associations were reported between short-term exposure to PM 10-2.5 andhospital admissions or ED visits for cardiovascular causes. These results are supported by a largeU.S. multicity study of older adults that reported PM 10-2.5 associations with CVD hospitaladmissions, and only a slight reduction in the PM 10-2.5 risk estimate when included in a copollutantmodel with PM 2.5 (Section 6.2.10). The PM 10-2.5 associations with cardiovascular hospital admissionsand ED visits were observed in study locations with mean 24-h avg PM 10-2.5 concentrations rangingfrom 7.4 to 13 μg/m3. These results are supported by the associations observed between PM 10-2.5 andcardiovascular mortality in areas with 24-h avg PM 10-2.5 concentrations ranging from 6.1-16.4 μg/m3(Section 6.2.11). The results of the epidemiologic studies were further confirmed by studies thatexamined dust storm events, which contain high concentrations of crustal material, and found anincrease in cardiovascular-related ED visits and hospital admissions. Additional epidemiologicstudies have reported PM 10-2.5 associations with other cardiovascular health effects includingsupraventricular ectopy and changes in HRV (Section 6.2.1.1). Although limited in number, studiesof controlled human exposures provide some evidence to support the alterations in HRV observed inthe epidemiologic studies (Section 6.2.1.2). The few toxicological studies that examined the effect ofPM 10-2.5 on cardiovascular health effects used IT instillation due to the technical challenges inexposing rodents via inhalation to PM 10-2.5 , and, as a result, provide only limited evidence on thebiological plausibility of PM 10-2.5 induced cardiovascular effects. The potential for PM 10-2.5 to elicitan effect is supported by dosimetry studies, which show that a large proportion of inhaled particles inthe 3-6 micron (d ae ) range can reach and deposit in the lower respiratory tract, particularly thetracheobronchial (TB) airways (Figures 4-3 and 4-4). Collectively, the evidence from epidemiologicstudies, along with the more limited evidence from controlled human exposure and toxicologicalstudies is suggestive of a causal relationship between short-term exposures to PM 10-2.5and cardiovascular effects.

Respiratory Effects

A number of recent epidemiologic studies conducted in Canada and France found consistent,positive associations between respiratory ED visits and hospital admissions and short-term exposureto PM 10-2.5 in studies with mean 24-h avg concentrations ranging from 5.6-16.2 μg/m3 (Section 6.3.8) .In these studies, the strongest relationships were observed among children, with less consistentevidence for adults and older adults (i.e., ≥ 65). In a large multicity study of older adults, PM 10-2.5was positively associated with respiratory hospital admissions in both single and copollutant modelswith PM 2.5 . In addition, a U.S.-based multicity study found evidence for an increase in respiratorymortality upon short-term exposure to PM 10-2.5 , but these associations have not been consistentlyobserved in single-city studies (Section 6.3.9). A limited number of epidemiologic studies havefocused on specific respiratory morbidity outcomes, and found no evidence of an association withlower respiratory symptoms, wheeze, and medication use (Section 6.3.1.1). While controlled humanexposure studies have not observed an effect on lung function or respiratory symptoms in healthy orasthmatic adults in response to short-term exposure to PM 10-2.5 , healthy volunteers have exhibited anincrease in markers of pulmonary inflammation. Toxicological studies using inhalation exposures arestill lacking, but pulmonary injury has been observed in animals after IT instillation exposure(Section 6.3.5.3). In some cases, PM 10-2.5 was found to be more potent than PM 2.5 and effects werenot attributable to endotoxin. Both rural and urban PM 10-2.5 have induced inflammation and injuryresponses in rats or mice exposed via IT instillation, making it difficult to distinguish the healtheffects of PM 10-2.5 from different environments. Overall, epidemiologic studies, along with thelimited number of controlled human exposure and toxicological studies that examined PM 10-2.5respiratory effects provide evidence that is suggestive of a causal relationship between short-term exposures to PM 10-2.5 and respiratory effects.

Mortality

The majority of studies evaluated in this review provide some evidence for mortalityassociations with PM 10-2.5 in areas with mean 24-h avg concentrations ranging from 6.1-16.4 μg/m3.However, uncertainty surrounds the PM 10-2.5 associations reported in the studies evaluated due to thedifferent methods used to estimate PM 10-2.5 concentrations across studies (e.g., direct measurementof PM 10-2.5 using dichotomous samplers, calculating the difference between PM 10 and PM 2.5concentrations). In addition, only a limited number of PM 10-2.5 studies have investigated potentialconfounding by gaseous copollutants or the influence of model specification on PM 10-2.5 riskestimates.

A new U.S.-based multicity study, which estimated PM 10-2.5 concentrations by calculating thedifference between the county-average PM 10 and PM 2.5 , found associations between PM 10-2.5 andmortality across the U.S., including evidence for regional variability in PM 10-2.5 risk estimates(Section 6.5.2.3). Additionally, the U.S.-based multicity study provides preliminary evidence forgreater effects occurring during the warmer months (i.e., spring and summer). A multicity Canadianstudy provides additional evidence for an association between short-term exposure to PM 10-2.5 andmortality (Section 6.5.2.3). Although consistent positive associations have been observed across bothmulti- and single-city studies, more data are needed to adequately characterize the chemical andbiological components that may modify the potential toxicity of PM 10-2.5 and compare the differentmethods used to estimate exposure. Overall, the evidence evaluated is suggestive of a causalrelationship between short-term exposures to PM 10-2.5 and mortality.

2.3.4. Integration of PM 10-2.5 Effects

Epidemiologic, controlled human exposure, and toxicological studies have provided evidence that issuggestive for relationships between short-term exposure to PM 10-2.5 and cardiovascular effects,respiratory effects, and mortality. Conclusions regarding causation for the various health effects andoutcomes were made for PM 10-2.5 as a whole regardless of origin, since PM 10-2.5 -related effects havebeen demonstrated for a number of different environments (e.g., cities reflecting a wide range ofenvironmental conditions). Associations between short-term exposure to PM 10-2.5 and cardiovascularand respiratory effects, and mortality have been observed in locations with mean PM 10-2.5concentrations ranging from 5.6 to 33.2 μg/m3, and maximum PM 10-2.5 concentrations ranging from24.6 to 418.0 μg/m3) (Figure 2-3). A number of different health effects are included in Figure 2-3 toprovide an integration of the range of effects by mean concentration, with a focus on cardiovascularand respiratory effects, and mortality (i.e., health effects categories with at least a suggestive causaldetermination). To date, a sufficient amount of evidence does not exist in order to draw conclusionsregarding the health effects and outcomes associated with long-term exposure to PM 10-2.5 .

In epidemiologic studies, associations between short-term exposure to PM 10-2.5 andcardiovascular outcomes (i.e., IHD hospital admissions, supraventricular ectopy, and changes inHRV) have been found that are similar in magnitude to those observed in PM 2.5 studies. Controlledhuman exposure studies have also observed alterations in HRV, providing consistency and coherencefor the effects observed in the epidemiologic studies. To date, only a limited number of toxicologicalstudies have been conducted to examine the effects of PM 10-2.5 on cardiovascular effects. All of thesestudies involved IT instillation due to the technical challenges of using PM 10-2.5 for rodent inhalationstudies. As a result, the toxicological studies evaluated provide limited biological plausibility for thePM 10-2.5 effects observed in the epidemiologic and controlled human exposure studies.

Figure 2-3. Summary of U.S. studies examining the association between short-term exposure to PM 10-2.5 and cardiovascular morbidity/mortality and respiratory morbidity/mortality. All effect estimates have been standardized to reflect a10 μg/m3 increase in mean 24-h avg PM 10-2.5 concentration and ordered by increasing concentration.

Limited evidence is available from epidemiologic studies for respiratory health effects andoutcomes in response to short-term exposure to PM 10-2.5 . An increase in respiratory hospitaladmissions and ED visits has been observed, but primarily in studies conducted in Canada andEurope. In addition, associations are not reported for lower respiratory symptoms, wheeze, ormedication use. Controlled human exposure studies have not observed an effect on lung function orrespiratory symptoms in healthy or asthmatic adults, but healthy volunteers have exhibitedpulmonary inflammation. The toxicological studies (all IT instillation) provide evidence ofpulmonary injury and inflammation. In some cases, PM 10-2.5 was found to be more potent than PM 2.5and effects were not solely attributable to endotoxin.

Currently, a national network is not in place to monitor PM 10-2.5 concentrations. As a result,uncertainties surround the concentration at which the observed associations occur. Ambientconcentrations of PM 10-2.5 are generally determined by the subtraction of PM 10 and PM 2.5measurements, using various methods. For example, some epidemiologic studies estimate PM 10-2.5by taking the difference between collocated PM 10 and PM 2.5 monitors while other studies have takenthe difference between county average PM 10 and PM 2.5 concentrations. Moreover, there are potentialdifferences among operational flow rates and temperatures for PM 10 and PM 2.5 monitors used tocalculate PM 10-2.5 . Therefore, there is greater error in ambient exposure to PM 10-2.5 compared toPM 2.5 . This would tend to increase uncertainty and make it more difficult to detect effects of PM 10-2.5in epidemiologic studies. In addition, the various differences between eastern and western U.S.counties can lead to exposure misclassification, and the potential underestimation of effects inwestern counties (as discussed for PM 2.5 in Section 2.3.2).

The 2004 PM AQCD presented the limited amount of evidence available that examined thepotential association between exposure to PM 10-2.5 and health effects and outcomes. The currentevidence, primarily from epidemiologic studies, builds upon the results from the 2004 PM AQCDand indicates that short-term exposure to PM 10-2.5 is associated with effects on both thecardiovascular and respiratory systems. However, variability in the chemical and biologicalcomposition of PM 10-2.5 , limited evidence regarding effects of the various components of PM 10-2.5 ,and lack of clearly defined biological mechanisms for PM 10-2.5 -related effects are important sourcesof uncertainty.

2.3.5. Exposure to UFPs

2.3.5.1. Effects of Short-Term Exposure to UFPs

Cardiovascular Effects

Controlled human exposure studies provide the majority of the evidence for cardiovascularhealth effects in response to short-term exposure to UFPs. While there are a limited number ofstudies that have examined the association between UFPs and cardiovascular morbidity, there is alarger body of evidence from studies that exposed subjects to fresh DE, which is typically dominatedby UFPs. These studies have consistently demonstrated changes in vasomotor function followingexposure to atmospheres containing relatively high concentrations of particles (Section 6.2.4.2).Markers of systemic oxidative stress have also been observed to increase after exposure to variousparticle types that are predominantly in the UFP size range. In addition, alterations in HRVparameters have been observed in response to controlled human exposure to UF CAPs, withinconsistent evidence for changes in markers of blood coagulation following exposure to UF CAPsand DE (Sections 6.2.1.2 and 6.2.8.2). A few toxicological studies have also found consistentchanges in vasomotor function, which provides coherence with the effects demonstrated in thecontrolled human exposure studies (Section 6.2.4.3). Additional UFP-induced effects observed intoxicological studies include alterations in HRV, with less consistent effects observed for systemicinflammation and blood coagulation. Only a few epidemiologic studies have examined the effect ofUFPs on cardiovascular morbidity and collectively they found inconsistent evidence for anassociation between UFPs and CVD hospital admissions, but some positive associations forsubclinical cardiovascular measures (i.e., arrhythmias and supraventricular beats) (Section 6.2.2.1).These studies were conducted in the U.S. and Europe in areas with mean particle numberconcentration ranging from ~8,500 to 36,000 particles/cm3. However, UFP number concentrationsare highly variable (i.e., concentrations drop off quickly from the road compared to accumulationmode particles), and therefore, more subject to exposure error than accumulation mode particles. Inconclusion, the evidence from the studies evaluated is suggestive of a causal relationshipbetween short-term exposures to UFPs and cardiovascular effects.

Respiratory Effects

A limited number of epidemiologic studies have examined the potential association betweenshort-term exposure to UFPs and respiratory morbidity. Of the studies evaluated, there is limited, andinconsistent evidence for an association between short-term exposure to UFPs and respiratorysymptoms, as well as asthma hospital admissions in locations a median particle numberconcentration of ~6,200 to a mean of 38,000 particles/cm3 (Section 6.3.10). The spatial and temporalvariability of UFPs also affects these associations. Toxicological studies have reported respiratoryeffects including oxidative, inflammatory, and allergic responses using a number of different UFPtypes (Section 6.3). Although controlled human exposure studies have not extensively examined theeffect of UFPs on respiratory outcomes, a few studies have observed small UFP-inducedasymptomatic decreases in pulmonary function. Markers of pulmonary inflammation have beenobserved to increase in healthy adults following controlled exposures to UFPs, particularly in studiesusing fresh DE. However, it is important to note that for both controlled human exposure and animaltoxicological studies of exposures to fresh DE, the relative contributions of gaseous copollutants tothe respiratory effects observed remain unresolved. Thus, the current collective evidence issuggestive of a causal relationship between short-term exposures to UFPs andrespiratory effects.

2.3.6. Integration of UFP Effects

The controlled human exposure studies evaluated have consistently demonstrated effects onvasomotor function and systemic oxidative stress with additional evidence for alterations in HRVparameters in response to exposure to UF CAPs. The toxicological studies provide coherence for thechanges in vasomotor function observed in the controlled human exposure studies. Epidemiologicstudies are limited because a national network is not in place to measure UFP in the U.S. UFPconcentrations are spatially and temporally variable, which would increase uncertainty and make itdifficult to detect associations between health effects and UFPs in epidemiologic studies. In addition,data on the composition of UFPs, the spatial and temporal evolution of UFP size distribution andchemical composition, and potential effects of UFP constituents are sparse.

More limited evidence is available regarding the effect of UFPs on respiratory effects.Controlled human exposure studies have not extensively examined the effect of UFPs on respiratorymeasurements, but a few studies have observed small decrements in pulmonary function andincreases in pulmonary inflammation. Additional effects including oxidative, inflammatory, and pro-allergic outcomes have been demonstrated in toxicological studies. Epidemiologic studies havefound limited and inconsistent evidence for associations between UFPs and respiratory effects.

Overall, a limited number of studies have examined the association between exposure to UFPsand morbidity and mortality. Of the studies evaluated, controlled human exposure and toxicologicalstudies provide the most evidence for UFP-induced cardiovascular and respiratory effects; however,many studies focus on exposure to DE. As a result, it is unclear if the effects observed are due toUFP, larger particles (i.e., PM 2.5 ), or the gaseous components of DE. Additionally, UF CAPs systemsare limited as the atmospheric UFP composition is modified when concentrated, which addsuncertainty to the health effects observed in controlled human exposure studies (Section 1.5.3).

2.4. Policy Relevant Considerations

2.4.1. Potentially Susceptible Populations

Upon evaluating the association between short- and long-term exposure to PM and varioushealth outcomes, studies also attempted to identify populations that are more susceptible to PM (i.e.,populations that have a greater likelihood of experiencing health effects related to exposure to an airpollutant (e.g., PM) due to a variety of factors including, but not limited to: genetic or developmentalfactors, race, gender, life stage, lifestyle (e.g., smoking status and nutrition) or preexisting disease; aswell as, population-level factors that can increase an individual's exposure to an air pollutant (e.g.,PM) such as socioeconomic status [SES], which encompasses reduced access to health care, loweducational attainment, residential location, and other factors). These studies did so by conductingstratified analyses; by examining effects in individuals with an underlying health condition; or bydeveloping animal models that mimic the pathophysiologic conditions associated with an adversehealth effect. In addition, numerous studies that focus on only one potentially susceptible populationprovide supporting evidence on whether a population is susceptible to PM exposure. These studiesidentified a multitude of factors that could potentially contribute to whether an individual issusceptible to PM (Table 8-2). Although studies have primarily used exposures to PM 2.5 or PM 10 , theavailable evidence suggests that the identified factors may also enhance susceptibility to PM 10-2.5 .The examination of susceptible populations to PM exposure allows for the NAAQS to provide anadequate margin of safety for both the general population and for susceptible populations.

During specific periods of life (i.e., childhood and advanced age), individuals may be moresusceptible to environmental exposures, which in turn can render them more susceptible to PM-related health effects. An evaluation of age-related health effects suggests that older adults haveheightened responses for cardiovascular morbidity with PM exposure. In addition, epidemiologicand toxicological studies provide evidence that indicates children are at an increased risk of PM-related respiratory effects. It should be noted that the health effects observed in children could beinitiated by exposures to PM that occurred during key windows of development, such as in utero.Epidemiologic studies that focus on exposures during development have reported inconsistentfindings (Section 7.4), but a recent toxicological study suggests that inflammatory responses inpregnant women due to exposure to PM could result in health effects in the developing fetus.

Epidemiologic studies have also examined whether additional factors, such as gender, race, orethnicity modify the association between PM and morbidity and mortality outcomes. Althoughgender and race do not seem to modify PM risk estimates, limited evidence from two studiesconducted in California suggest that Hispanic ethnicity may modify the association between PM andmortality.

Recent epidemiologic and toxicological studies provided evidence that individuals with nullalleles or polymorphisms in genes that mediate the antioxidant response to oxidative stress (i.e.,GSTM1), regulate enzyme activity (i.e., MTHFR and cSHMT), or regulate levels of procoagulants(i.e., fibrinogen) are more susceptible to PM exposure. However, some studies have shown thatpolymorphisms in genes (e.g., HFE) can have a protective effect against effects of PM exposure.Additionally, preliminary evidence suggests that PM exposure can impart epigenetic effects (i.e.,DNA methylation); however, this requires further investigation.

Recently studies have begun to examine the influence of preexisting chronic inflammatoryconditions, such as diabetes and obesity, on PM-related health effects. These studies have foundsome evidence for increased associations for cardiovascular outcomes along with pathophysiologicalterations in markers of inflammation, oxidative stress, and acute phase response. However, moreresearch is needed to thoroughly examine the affect of PM exposure on obese individuals and toidentify the biological pathway(s) that could increase the susceptibility of diabetic and obeseindividuals to PM.

There is also evidence that SES, measured using surrogates such as educational attainment orresidential location, modifies the association between PM and morbidity and mortality outcomes. Inaddition, nutritional status, another surrogate measure of SES, has been shown to have protectiveeffects against PM exposure in individuals that have a higher intake of some vitamins and nutrients.

Overall, the epidemiologic, controlled human exposure, and toxicological studies evaluated inthis review provide evidence for increased susceptibility for various populations, including childrenand older adults, people with pre-existing cardiopulmonary diseases, and people with lower SES.

2.4.2. Lag Structure of PM-Morbidity and PM-Mortality Associations

Epidemiologic studies have evaluated the time-frame in which exposure to PM can impart ahealth effect. PM exposure-response relationships can potentially be influenced by a multitude offactors, such as the underlying susceptibility of an individual (e.g., age, pre-existing diseases), whichcould increase or decrease the lag times observed.

An attempt has been made to identify whether certain lag periods are more strongly associatedwith specific health outcomes. The epidemiologic evidence evaluated in the 2004 PM AQCDsupported the use of lags of 0-1 days for cardiovascular effects and longer moving averages ordistributed lags for respiratory diseases (U.S. EPA, 2004, 056905). However, currently, littleconsensus exists as to the most appropriate a priori lag times to use when examining morbidity andmortality outcomes. As a result, many investigators have chosen to examine the lag structure ofassociations between PM concentration and health outcome instead of focusing on a priori lag times.This approach is informative because if effects are cumulative, higher overall risks may exist thanwould be observed for any given single-day lag.

2.4.2.1. PM-Cardiovascular Morbidity Associations

Most of the studies evaluated that examined the association between cardiovascular hospitaladmissions and ED visits report associations with short-term PM exposure at lags 0- to 2-days, withmore limited evidence for shorter durations (i.e., hours) between exposure and response for somehealth effects (e.g., onset of MI) (Section 6.2.10). However, these studies have rarely examinedalternative lag structures. Controlled human exposure and toxicological studies provide biologicalplausibility for the health effects observed in the epidemiologic studies at immediate or concurrentday lags. Although the majority of the evidence supports shorter lag times for cardiovascular healtheffects, a recent study has provided preliminary evidence suggesting that longer lag times (i.e., 14-day distributed lag model) may be plausible for non-ischemic cardiovascular conditions(Section 6.2.10). Panel studies of short-term exposure to PM and cardiovascular endpoints have alsoexamined the time frame from exposure to health effect using a wide range of lag times. Studies ofECG changes indicating ischemia show effects at lags from several hours to 2 days, while lag timesranging from hours to several week moving averages have been observed in studies of arrhythmias,vasomotor function and blood markers of inflammation, coagulation and oxidative stress(Section 6.2). The longer lags observed in these panel studies may be explained if the effects of PMare cumulative. Although few studies of cumulative effects have been conducted, toxicologicalstudies have demonstrated PM-dependent progression of atherosclerosis. It should be noted that PMexposure could also lead to an acute event (e.g., infarction or stroke) in individuals withatherosclerosis that may have progressed in response to cumulative PM exposure. Therefore, effectshave been observed at a range of lag periods from a few hours to several days with no clear evidencefor any lag period having stronger associations then another.

2.4.2.2. PM-Respiratory Morbidity Associations

Generally, recent studies of respiratory hospital admissions that evaluate multiple lags, havefound effect sizes to be larger when using longer moving averages or distributed lag models. Forexample, when examining hospital admissions for all respiratory diseases among older adults, thestrongest associations were observed when using PM concentrations 2 days prior to the hospitaladmission (Section 6.3.8). Longer lag periods were also found to be most strongly associated withasthma hospital admissions and ED visits in children (3-5 days) with some evidence for moreimmediate effects in older adults (lags of 0 and 1 day), but these observations were not consistentacross studies (Section 6.3.8). These variable results could be due to the biological complexity ofasthma, which inhibits the identification of a specific lag period. The longer lag times identified inthe epidemiologic studies evaluated are biologically plausible considering that PM effects on allergicsensitization and lung immune defenses have been observed in controlled human exposure andtoxicological studies. These effects could lead to respiratory illnesses over a longer time course (e.g.,within several days respiratory infection may become evident, resulting in respiratory symptoms or ahospital admission). However, inflammatory responses, which contribute to some forms of asthma,may result in symptoms requiring medical care within a shorter time frame (e.g., 0-1 days).

2.4.2.3. PM-Mortality Associations

Epidemiologic studies that focused on the association between short-term PM exposure andmortality (i.e., all-cause, cardiovascular, and respiratory) mostly examined a priori lag structures ofeither 1 or 0-1 days. Although mortality studies do not often examine alternative lag structures, theselection of the aforementioned a priori lag days has been confirmed in additional studies, with thestrongest PM-mortality associations consistently being observed at lag 1 and 0-1-days (Section 6.5).However, of note is recent evidence for larger effect estimates when using a distributed lag model.

Epidemiologic studies that examined the association between long-term exposure to PM andmortality have also attempted to identify the latency period from PM exposure to death(Section 7.6.4). Results of the lag comparisons from several cohort studies indicate that the effects ofchanges in exposure on mortality are seen within five years, with the strongest evidence for effectsobserved within the first two years. Additionally, there is evidence, albeit from one study, that themortality effect had larger cumulative effects spread over the follow-up year and three precedingyears.

2.4.3. PM Concentration-Response Relationship

An important consideration in characterizing the PM-morbidity and mortality association iswhether the concentration-response relationship is linear across the full concentration range that isencountered or if there are concentration ranges where there are departures from linearity(i.e., nonlinearity). In this ISA studies have been identified that attempt to characterize the shape ofthe concentration-response curve along with possible PM “thresholds” (i.e., levels which PMconcentrations must exceed in order to elicit a health response). The epidemiologic studies evaluatedthat examined the shape of the concentration-response curve and the potential presence of athreshold have focused on cardiovascular hospital admissions and ED visits and mortality associatedwith short-term exposure to PM 10 and mortality associated with long-term exposure to PM 2.5 .

A limited number of studies have been identified that examined the shape of the PM-cardiovascular hospital admission and ED visit concentration-response relationship. Of these studies,some conducted an exploratory analysis during model selection to determine if a linear curve mostadequately represented the concentration-response relationship; whereas, only one study conductedan extensive analysis to examine the shape of the concentration-response curve at differentconcentrations (Section 6.2.10.10). Overall, the limited evidence from the studies evaluated supportsthe use of a no-threshold, log-linear model, which is consistent with the observations made in studiesthat examined the PM-mortality relationship.

Although multiple studies have previously examined the PM-mortality concentration-responserelationship and whether a threshold exists, more complex statistical analyses continue to bedeveloped to analyze this association. Using a variety of methods and models, most of the studiesevaluated support the use of a no-threshold, log-linear model; however, one study did observeheterogeneity in the shape of the concentration-response curve across cities (Section 6.5). Overall,the studies evaluated further support the use of a no-threshold log-linear model, but additional issuessuch as the influence of heterogeneity in estimates between cities, and the effect of seasonal andregional differences in PM on the concentration-response relationship still require furtherinvestigation.

In addition to examining the concentration-response relationship between short-term exposureto PM and mortality, Schwartz et al. (2008, 156963) conducted an analysis of the shape of theconcentration-response relationship associated with long-term exposure to PM. Using a variety ofstatistical methods, the concentration-response curve was found to be indistinguishable from linear,and, therefore, little evidence was observed to suggest that a threshold exists in the associationbetween long-term exposure to PM 2.5 and the risk of death (Section 7.6)

2.4.4. PM Sources and Constituents Linked to Health Effects

Recent epidemiologic, toxicological, and controlled human exposure studies have evaluatedthe health effects associated with ambient PM constituents and sources, using a variety ofquantitative methods applied to a broad set of PM constituents, rather than selecting a fewconstituents a priori (Section 6.6). There is some evidence for trends and patterns that link particularambient PM constituents or sources with specific health outcomes, but there is insufficient evidenceto determine whether these patterns are consistent or robust.

For cardiovascular effects, multiple outcomes have been linked to a PM 2.5 crustal/soil/roaddust source, including cardiovascular mortality and ST-segment changes. Additional studies havereported associations between other sources (i.e., traffic and wood smoke/vegetative burning) andcardiovascular outcomes (i.e., mortality and ED visits). Studies that only examined the effects ofindividual PM 2.5 constituents found evidence for an association between EC and cardiovascularhospital admissions and cardiovascular mortality. Many studies have also observed associationsbetween other sources (i.e., salt, secondary SO 4 2–/long-range transport, other metals) andcardiovascular effects, but at this time, there does not appear to be a consistent trend or pattern ofeffects for those factors.

There is less consistent evidence for associations between PM sources and respiratory healtheffects, which may be partially due to the fact that fewer source apportionment studies have beenconducted that examined respiratory-related outcomes (e.g., hospital admissions) and measures (e.g.,lung function). However, there is some evidence for associations between respiratory ED visits anddecrements in lung function with secondary SO 4 2– PM 2.5 . In addition, crustal/soil/road dust andtraffic sources of PM have been found to be associated with increased respiratory symptoms inasthmatic children and decreased PEF in asthmatic adults. Inconsistent results were observed inthose PM 2.5 studies that used individual constituents to examine associations with respiratorymorbidity and mortality, although Cu, Pb, OC, and Zn were related to respiratory health effects intwo or more studies.

A few studies have identified PM 2.5 sources associated with total mortality. These studiesfound an association between mortality and the PM 2.5 sources: secondary SO 4 2–/long-rangetransport, traffic, and salt. In addition, studies have evaluated whether the variation in associationsbetween PM 2.5 and mortality or PM 10 and mortality reflects differences in PM 2.5 constituents. PM 10 -mortality effect estimates were greater in areas with a higher proportion of Ni in PM 2.5 , but theoverall PM 10 -mortality association was diminished when New York City was excluded in sensitivityanalyses in two of the studies. V was also found to modify PM 10 -mortality effect estimates. Whenexamining the effect of species-to-PM 2.5 mass proportion on PM 2.5 -mortality effect estimates, Ni,but not V, was also found to modify the association.

Overall, the results indicate that many constituents of PM can be linked with differing healtheffects and the evidence is not yet sufficient to allow differentiation of those constituents or sourcesthat are more closely related to specific health outcomes. These findings are consistent with theconclusions of the 2004 PM AQCD (U.S. EPA, 2004, 056905) (i.e., that a number of source types,including motor vehicle emissions, coal combustion, oil burning, and vegetative burning, areassociated with health effects). Although the crustal factor of fine particles was not associated withmortality in the 2004 PM AQCD (U.S. EPA, 2004, 056905), recent studies have suggested that PM(both PM 2.5 and PM 10-2.5 ) from crustal, soil or road dust sources or PM tracers linked to these sourcesare associated with cardiovascular effects. In addition, PM 2.5 secondary SO 4 2– has been associatedwith both cardiovascular and respiratory effects.

2.5. Welfare Effects

This section presents key conclusions and scientific judgments regarding causality for welfareeffects of PM as discussed in Chapter 9. The effects of particulate NO X and SO X have recently beenevaluated in the ISA for Oxides of Nitrogen and Sulfur – Ecological Criteria (U.S. EPA, 2008,157074). That ISA focused on the effects from deposition of gas- and particle-phase pollutantsrelated to ambient NO X and SO X concentrations that can lead to acidification and nutrientenrichment. Thus, emphasis in Chapter 9 is placed on the effects of airborne PM, including NO X andSO X , on visibility and climate, and on the effects of deposition of PM constituents other than NO Xand SO X , primarily metals and carbonaceous compounds. EPA’s framework for causality, describedin Chapter 1, was applied and the causal determinations are highlighted.

2.5.1. Summary of Effects on Visibility

Visibility impairment is caused by light scattering and absorption by suspended particles andgases. There is strong and consistent evidence that PM is the overwhelming source of visibilityimpairment in both urban and remote areas. EC and some crustal minerals are the only commonlyoccurring airborne particle components that absorb light. All particles scatter light, and generallylight scattering by particles is the largest of the four light extinction components (i.e., absorption andscattering by gases and particles). Although a larger particle scatters more light than a similarlyshaped smaller particle of the same composition, the light scattered per unit of mass is greatest forparticles with diameters from ~0.3-1.0 μm.

For studies where detailed data on particle composition by size are available, accuratecalculations of light extinction can be made. However, routinely available PM speciation data can beused to make reasonable estimates of light extinction using relatively simple algorithms that multiplythe concentrations of each of the major PM species by its dry extinction efficiency and by a watergrowth term that accounts for particle size change as a function of relative humidity for hygroscopicspecies (e.g., sulfate, nitrate, and sea salt). This permits the visibility impairment associated witheach of the major PM components to be separately approximated from PM speciation monitoringdata.

Direct optical measurement of light extinction measured by transmissometer, or by combiningthe PM light scattering measured by integrating nephelometers with the PM light absorptionmeasured by an aethalometer, offer a number of advantages compared to algorithm estimates of lightextinction based on PM composition and relative humidity data. The direct measurements are notsubject to the uncertainties associated with assumed scattering and absorption efficiencies used in thePM algorithm approach. The direct measurements have higher time resolution (i.e., minutes tohours), which is more commensurate with visibility effects compared with calculated light extinctionusing routinely available PM speciation data (i.e., 24-h duration).

Particulate sulfate and nitrate have comparable light extinction efficiencies (haze impacts perunit mass concentration) at any relative humidity value. Their light scattering per unit massconcentration increases with increasing relative humidity, and at sufficiently high humidity values(RH>85%) they are the most efficient particulate species contributing to haze. Particulate sulfate isthe dominant source of regional haze in the eastern U.S. (>50% of the particulate light extinction)and an important contributor to haze elsewhere in the country (>20% of particulate light extinction).Particulate nitrate is a minor component of remote-area regional haze in the non-California westernand eastern U.S., but an important contributor in much of California and in the upper MidwesternU.S., especially during winter when it is the dominant contributor to particulate light extinction.

EC and OC have the highest dry extinction efficiencies of the major PM species and areresponsible for a large fraction of the haze, especially in the northwestern U.S., though absoluteconcentrations are as high in the eastern U.S. Smoke plume impacts from large wildfires dominatemany of the worst haze periods in the western U.S. Carbonaceous PM is generally the largestcomponent of urban excess PM 2.5 (i.e., the difference between urban and regional backgroundconcentration). Western urban areas have more than twice the average concentrations ofcarbonaceous PM than remote areas sites in the same region. In eastern urban areas PM 2.5 isdominated by about equal concentrations of carbonaceous and sulfate components, though theusually high relative humidity in the East causes the hydrated sulfate particles to be responsible forabout twice as much of the urban haze as that caused by the carbonaceous PM.

PM 2.5 crustal material (referred to as fine soil) and PM 10-2.5 are significant contributors to hazefor remote areas sites in the arid southwestern U.S. where they contribute a quarter to a third of thehaze, with PM 10-2.5 usually contributing twice that of fine soil. Coarse mass concentrations are ashigh in the Central Great Plains as in the deserts though there are no corresponding highconcentrations of fine soil as in the Southwest. Also the relative contribution to haze by the highcoarse mass in the Great Plains is much smaller because of the generally higher haze values causedby the high concentrations of sulfate and nitrate PM in that region.

Visibility has direct significance to people’s enjoyment of daily activities and their overallsense of wellbeing. For example, psychological research has demonstrated that people areemotionally affected by poor VAQ such that their overall sense of wellbeing is diminished. Urbanvisibility has been examined in two types of studies directly relevant to the NAAQS review process:urban visibility preference studies and urban visibility valuation studies. Both types of studies aredesigned to evaluate individuals’ desire for good VAQ where they live, using different metrics.Urban visibility preference studies examine individuals’ preferences by investigating the amount ofvisibility degradation considered unacceptable, while economic studies examine the value anindividual places on improving VAQ by eliciting how much the individual would be willing to payfor different amounts of VAQ improvement.

There are three urban visibility preference studies and two additional pilot studies that havebeen conducted to date that provide useful information on individuals’ preferences for good VAQ inthe urban setting. The completed studies were conducted in Denver, Colorado, two cities in BritishColumbia, Canada, and Phoenix, AZ. The additional studies were conducted in Washington, DC. Therange of median preference values for an acceptable amount of visibility degradation from the 4urban areas was approximately 19-33 dv. Measured in terms of visual range (VR), these medianacceptable values were between approximately 59 and 20 km.

The economic importance of urban visibility has been examined by a number of studiesdesigned to quantify the benefits (or willingness to pay) associated with potential improvements inurban visibility. Urban visibility valuation research was described in the 2004 PM AQCD (U.S. EPA,2004, 056905) and the 2005 PM Staff Paper (U.S. EPA, 2005, 090209). Since the mid-1990s, littlenew information has become available regarding urban visibility valuation (Section 9.2.4).

Collectively, the evidence is sufficient to conclude that a causal relationship existsbetween PM and visibility impairment.

2.5.2. Summary of Effects on Climate

Aerosols affect climate through direct and indirect effects. The direct effect is primarilyrealized as planet brightening when seen from space because most aerosols scatter most of thevisible spectrum light that reaches them. The Intergovernmental Panel on Climate Change (IPCC)Fourth Assessment Report (AR4) (IPCC, 2007, 092765), hereafter IPCC AR4, reported that theradiative forcing from this direct effect was -0.5 (±0.4) W/m2 and identified the level of scientificunderstanding of this effect as 'Medium-low'. The global mean direct radiative forcing effect fromindividual components of aerosols was estimated for the first time in the IPCC AR4 where they werereported to be (all in W/m2 units): -0.4 (±0.2) for sulfate, -0.05 (±0.05) for fossil fuel-derived organiccarbon, +0.2 (±0.15) for fossil fuel-derived black carbon (BC), +0.03 (±0.12) for biomass burning,-0.1 (±0.1) for nitrates, and -0.1 (±0.2) for mineral dust. Global loadings of anthropogenic dust andnitrates remain very troublesome to estimate, making the radiative forcing estimates for theseconstituents particularly uncertain.

Numerical modeling of aerosol effects on climate has sustained remarkable progress since thetime of the 2004 PM AQCD (U.S. EPA, 2004, 056905), PM AQCD, though model solutions stilldisplay large heterogeneity in their estimates of the direct radiative forcing effect fromanthropogenic aerosols. The clear-sky direct radiative forcing over ocean due to anthropogenicaerosols is estimated from satellite instruments to be on the order of -1.1 (±0.37) W/m2 while modelestimates are -0.6 W/m2. The models' low bias over ocean is carried through for the global average:global average direct radiative forcing from anthropogenic aerosols is estimated from measurementsto range from -0.9 to -1.9 W/m2, larger than the estimate of -0.8 W/m2 from the models.

Aerosol indirect effects on climate are primarily realized as an increase in cloud brightness(termed the 'first indirect' or Twomey effect), changes in precipitation, and possible changes in cloudlifetime. The IPCC AR4 reported that the radiative forcing from the Twomey effect was -0.7 (range:-1.1 to +4) and identified the level of scientific understanding of this effect as “Low” in part owingto the very large unknowns concerning aerosol size distributions and important interactions withclouds. Other indirect effects from aerosols are not considered to be radiative forcing.

Taken together, direct and indirect effects from aerosols increase Earth's shortwave albedo orreflectance thereby reducing the radiative flux reaching the surface from the Sun. This produces netclimate cooling from aerosols. The current scientific consensus reported by IPCC AR4 is that thedirect and indirect radiative forcing from anthropogenic aerosols computed at the top of theatmosphere, on a global average, is about -1.3 (range: -2.2 to -0.5) W/m2. While the overall globalaverage effect of aerosols at the top of the atmosphere and at the surface is negative, absorption andscattering by aerosols within the atmospheric column warms the atmosphere between the Earth'ssurface and top of the atmosphere. In part, this is owing to differences in the distribution of aerosoltype and size within the vertical atmospheric column since aerosol type and size distributionsstrongly affect the aerosol scattering and reradiation efficiencies at different altitudes andatmospheric temperatures. And, although the magnitude of the overall negative radiative forcing atthe top of the atmosphere appears large in comparison to the analogous IPCC AR4 estimate ofpositive radiative forcing from anthropogenic GHG of about +2.9 (± 0.3) W/m 2, the horizontal,vertical, and temporal distributions and the physical lifetimes of these two very different radiativeforcing agents are not similar; therefore, the effects do not simply off-set one another.

Overall, the evidence is sufficient to conclude that a causal relationship exists betweenPM and effects on climate, including both direct effects on radiative forcing and indirecteffects that involve cloud feedbacks that influence precipitation formation and cloudlifetimes.

2.5.3. Summary of Ecological Effects of PM

Ecological effects of PM include direct effects to metabolic processes of plant foliage;contribution to total metal loading resulting in alteration of soil biogeochemistry and microbiology,plant growth and animal growth and reproduction; and contribution to total organics loadingresulting in bioaccumulation and biomagnification across trophic levels. These effects were well-characterized in the 2004 PM AQCD (U.S. EPA, 2004, 056905). Thus, the summary below buildsupon the conclusions provided in that review.

PM deposition comprises a heterogeneous mixture of particles differing in origin, size, andchemical composition. Exposure to a given concentration of PM may, depending on the mix ofdeposited particles, lead to a variety of phytotoxic responses and ecosystem effects. Moreover, manyof the ecological effects of PM are due to the chemical constituents (e.g., metals, organics, and ions)and their contribution to total loading within an ecosystem.

Investigations of the direct effects of PM deposition on foliage have suggested little or noeffects on foliar processes, unless deposition levels were higher than is typically found in theambient environment. However, consistent and coherent evidence of direct effects of PM has beenfound in heavily polluted areas adjacent to industrial point sources such as limestone quarries,cement kilns, and metal smelters (Sections 9.4.3 and 9.4.5.7). Where toxic responses have beendocumented, they generally have been associated with the acidity, trace metal content, surfactantproperties, or salinity of the deposited materials.

An important characteristic of fine particles is their ability to affect the flux of solar radiationpassing through the atmosphere, which can be considered in both its direct and diffuse components.Foliar interception by canopy elements occurs for both up- and down-welling radiation. Therefore,the effect of atmospheric PM on atmospheric turbidity influences canopy processes both by radiationattenuation and by changing the efficiency of radiation interception in the canopy throughconversion of direct to diffuse radiation. Crop yields can be sensitive to the amount of radiationreceived, and crop losses have been attributed to increased regional haze in some areas of the worldsuch as China (Section 9.4.4). On the other hand, diffuse radiation is more uniformly distributedthroughout the canopy and may increase canopy photosynthetic productivity by distributing radiationto lower leaves. The enrichment in photosynthetically active radiation (PAR) present in diffuseradiation may offset a portion of the effect of an increased atmospheric albedo due to atmosphericparticles. Further research is needed to determine the effects of PM alteration of radiative flux on thegrowth of vegetation in the U.S.

The deposition of PM onto vegetation and soil, depending on its chemical composition, canproduce responses within an ecosystem. The ecosystem response to pollutant deposition is a directfunction of the level of sensitivity of the ecosystem and its ability to ameliorate resulting change.Many of the most important ecosystem effects of PM deposition occur in the soil. Upon entering thesoil environment, PM pollutants can alter ecological processes of energy flow and nutrient cycling,inhibit nutrient uptake, change ecosystem structure, and affect ecosystem biodiversity. The soilenvironment is one of the most dynamic sites of biological interaction in nature. It is inhabited bymicrobial communities of bacteria, fungi, and actinomycetes, in addition to plant roots and soilmacro-fauna. These organisms are essential participants in the nutrient cycles that make elementsavailable for plant uptake. Changes in the soil environment can be important in determining plantand ultimately ecosystem response to PM inputs.

There is strong and consistent evidence from field and laboratory experiments that metalcomponents of PM alter numerous aspects of ecosystem structure and function. Changes in the soilchemistry, microbial communities and nutrient cycling, can result from the deposition of tracemetals. Exposures to trace metals are highly variable, depending on whether deposition is by wet ordry processes. Although metals can cause phytotoxicity at high concentrations, few heavy metals(e.g., Cu, Ni, Zn) have been documented to cause direct phytotoxicity under field conditions.Exposure to coarse particles and elements such as Fe and Mg are more likely to occur via drydeposition, while fine particles, which are more often deposited by wet deposition, are more likely tocontain elements such as Ca, Cr, Pb, Ni, and V. Ecosystems immediately downwind of majoremissions sources can receive locally heavy deposition inputs. Phytochelatins produced by plants asa response to sublethal concentrations of heavy metals are indicators of metal stress to plants.Increased concentrations of phytochelatins across regions and at greater elevation have beenassociated with increased amounts of forest injury in the northeastern U.S.

Overall, the ecological evidence is sufficient to conclude that a causal relationship is likelyto exist between deposition of PM and a variety of effects on individual organisms andecosystems, based on information from the previous review and limited new findings inthis review. However, in many cases, it is difficult to characterize the nature and magnitude ofeffects and to quantify relationships between ambient concentrations of PM and ecosystem responsedue to significant data gaps and uncertainties as well as considerable variability that exists in thecomponents of PM and their various ecological effects.

A significant detrimental effect of particle pollution is the soiling of painted surfaces and otherbuilding materials. Soiling changes the reflectance of opaque materials and reduces the transmissionof light through transparent materials. Soiling is a degradation process that requires remediation bycleaning or washing, and, depending on the soiled surface, repainting. Particulate deposition canresult in increased cleaning frequency of the exposed surface and may reduce the usefulness of thesoiled material.

Attempts have been made to quantify the pollutant exposure levels at which materials damageand soiling have been perceived. However, to date, insufficient data are available to advance theknowledge regarding perception thresholds with respect to pollutant concentration, particle size, andchemical composition. Nevertheless, the evidence is sufficient to conclude that a causalrelationship exists between PM and effects on materials.

2.6. Summary of Health Effects and Welfare Effects

Causal Determinations

This chapter has provided an overview of the underlying evidence used in making the causal determinations for the health and welfare effects and PM size fractions evaluated. This review builds upon the main conclusions of the last PM AQCD (U.S. EPA, 2004, 056905):

"A growing body of evidence both from epidemiological and toxicological studies supports the general conclusion that PM 2.5 (or one or more PM 2.5 components), acting alone and/or in

“A much more limited body of evidence is suggestive of associations between short-term (but not long-term) exposures to ambient coarse-fraction thoracic particles and various mortality and morbidity effects observed at times in some locations. This suggests that PM 10-2.5 , or some constituent component(s) of PM 10-2.5 , may contribute under some circumstances to increased human health risks with somewhat stronger evidence for associations with morbidity (especially respiratory) endpoints than for mortality.” (pg 9-79 and 9-80)

"Impairment of visibility in rural and urban areas is directly related to ambient concentrations of fine particles, as modulated by particle composition, size, and hygroscopic characteristics, and by relative humidity.” (pg 9-99)

“Available evidence, ranging from satellite to in situ measurements of aerosol effects on incoming solar radiation and cloud properties, is strongly indicative of an important role in climate for aerosols, but this role is still poorly quantified.” (pg 9-111)

The evaluation of the epidemiologic, toxicological, and controlled human exposure studies published since the completion of the 2004 PM AQCD have provided additional evidence for PM-related health effects. Table 2-6 provides an overview of the causal determinations for all PM size fractions and health effects. Causal determinations for PM and welfare effects, including visibility, climate, ecological effects, and materials are included in Table 2-7. Detailed discussions of the scientific evidence and rationale for these causal determinations are provided in the subsequent chapters of this ISA.

I noticed the monitor data has been way down since Sunday (although at least one downtown monitor seems to always be out of service) - the weather has much to do with pollution being washed and blown from your window so the last two days should have been better. Also, when I drove by Mittal today it seems like they have cut back on the polluting - keep an eye on their stacks...

I don;t know of any specific dangerous emissions from Air Products but I am checking into all the pollution inventory reports and all the monitors... even the PM 10 monitors they say we don't need any more... we'll soon know more about the air we are breathing.